the wetlands handbook

THE WETLANDS HANDBOOK

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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The Wetlands Handbook ED I TO RS Edward Maltby BSc PhD Professor of Wetland and Water Science Institute for Sustainable Water, Integrated Management and Ecosystem Research University of Liverpool Liverpool, L69 3GP, UK

Tom Barker BSc PhD Research Ecologist Institute for Sustainable Water, Integrated Management and Ecosystem Research University of Liverpool Liverpool, L69 3GP, UK

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This edition first published 2009, © 2009 by Blackwell Publishing Ltd Blackwell Publishing was acquired by John Wiley & Sons in February 2007. Blackwell’s publishing program has been merged with Wiley’s global Scientific, Technical and Medical business to form Wiley-Blackwell. Registered office: John Wiley & Sons Ltd, The Atrium, Southern Gate, Chichester, West Sussex, PO19 8SQ, UK Editorial offices: 9600 Garsington Road, Oxford, OX4 2DQ, UK The Atrium, Southern Gate, Chichester, West Sussex, PO19 8SQ, UK 111 River Street, Hoboken, NJ 07030-5774, USA For details of our global editorial offices, for customer services and for information about how to apply for permission to reuse the copyright material in this book please see our website at www.wiley.com/wiley-blackwell The right of the author to be identified as the author of this work has been asserted in accordance with the Copyright, Designs and Patents Act 1988. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, except as permitted by the UK Copyright, Designs and Patents Act 1988, without the prior permission of the publisher. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print may not be available in electronic books. Designations used by companies to distinguish their products are often claimed as trademarks. All brand names and product names used in this book are trade names, service marks, trademarks or registered trademarks of their respective owners. The publisher is not associated with any product or vendor mentioned in this book. This publication is designed to provide accurate and authoritative information in regard to the subject matter covered. It is sold on the understanding that the publisher is not engaged in rendering professional services. If professional advice or other expert assistance is required, the services of a competent professional should be sought. Library of Congress Cataloguing-in-Publication Data The wetlands handbook / edited by Edward Maltby, Tom Barker. p. cm. Includes bibliographical references. ISBN 978-0-632-05255-4 (hardback : alk. paper) 1. Wetlands. 2. Wetland management. I. Maltby, Edward. II. Barker, Tom. QH87.3.W479 2009 577.68–dc22 2008029043 A catalogue record for this book is available from the British Library. Set in 9/11.5 pt Trump Mediaeval by Newgen Imaging Systems (P) Ltd, Chennai, India Printed and bound in Singapore 1

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Contents

Preface, ix Contributors, xi SECTION I

WETLANDS IN THE GLOBAL ENVIRONMENT, 1

1 The Changing Wetland Paradigm, 3 Edward Maltby 2 Global Distribution, Diversity and Human Alterations of Wetland Resources, 43 Dennis F. Whigham 3

Biodiversity in Wetlands, 65 Brij Gopal

4

Peat as an Archive of Atmospheric, Climatic and Environmental Conditions, 96 R. Kelman Wieder, Merritt R. Turetsky and Melanie A. Vile

SECTION II WETLANDS IN THE NATURAL ENVIRONMENT: HOW DO WETLANDS WORK?, 113 5

Introduction – The Dynamics of Wetlands, 115 Tom Barker and Edward Maltby

6

Hydrological Dynamics I: Surface Waters, Flood and Sediment Dynamics, 120 Chris Baker, Julian R. Thompson and Matthew Simpson

7

Hydrological Dynamics II: Groundwater and Hydrological Connectivity, 169 Dave J. Gilvear and Chris Bradley

8

Hydrological Dynamics III: Hydro-ecology, 194 Ab P. Grootjans and Rudy Van Diggelen

9

Biogeochemical Dynamics I: Nitrogen Cycling in Wetlands, 213 John R. White and K.R. Reddy

10

Biogeochemical Dynamics II: Cycling and Storage of Phosphorus in Wetlands, 228 Curtis J. Richardson and Panchabi Vaithiyanathan

11

Biogeochemical Dynamics III: The Critical Role of Carbon in Wetlands, 249 Nancy B. Dise

12 Wetland Biogeochemical Cycles and their Interactions, 266 Jos T.A. Verhoeven

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Contents

13 Ecological Dynamics I: Vegetation as Bioindicator and Dynamic Community, 282 Bernard Clément and Michael C.F. Proctor 14 Ecological Dynamics II: The Influences of Vertebrate Herbivory on Ecological Dynamics in Wetland Ecosystems, 304 Isabel J.J. Van Den Wyngaert and Roland Bobbink 15 Ecological Dynamics III: Decomposition in Wetlands, 326 Scott D. Bridgham and Gary A. Lamberti SECTION III WETLANDS IN THE HUMAN ENVIRONMENT: HOW CAN WE UTILISE THE WORK OF WETLANDS?, 347 16 Introduction – Using Wetland Functioning, 349 Tom Barker and Edward Maltby 17 Wetlands and Water Resources, 357 Matthew P. McCartney and Michael C. Acreman 18 Wetland and Floodplain Soils: Their Characteristics, Management and Future, 382 Hadrian F. Cook, Samuel A.F. Bonnett and Leendert J. Pons 19 The Role of Buffer Zones for Agricultural Runoff, 417 Martin S.A. Blackwell, David V. Hogan, Gilles Pinay and Edward Maltby 20 Wetlands for Contaminant and Wastewater Treatment, 440 Robert H. Kadlec SECTION IV WETLAND ASSESSMENT: HOW CAN WE MEASURE THAT WETLANDS ARE WORKING?, 465 21 Introduction – Methodologies for Wetland Assessment, 467 Joseph S. Larson 22 The United States HGM (Hydrogeomorphic) Approach, 486 Mark M. Brinson 23 Development of a European Methodology for the Functional Assessment of Wetlands, 513 Edward Maltby, Tom Barker and Conor Linstead 24 Wetlands Assessment in Practice: Development and Application in the United States Regulatory Context, 545 R. Daniel Smith 25 Wetland Evaluation in Developing Countries, 569 Henri Roggeri 26

Methodologies for Economic Evaluation of Wetlands and Wetland Functioning, 601 R. Kerry Turner, Roy Brouwer and S. Georgiou

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Contents

vii

SECTION V WETLAND DYSFUNCTIONING: WHAT HAPPENS WHEN WETLANDS DO NOT WORK?, 627 27 Introduction – How Do Wetlands Fail?, 629 Katherine C. Ewel 28 Hydrological Impacts in and around Wetlands, 643 Michael C. Acreman and Matthew P. McCartney 29 Biotic Pressures and Their Effects on Wetland Functioning, 667 C. Max Finlayson 30 Human Impacts: Farming, Fire, Forestry and Fuel, 689 Hans Joosten SECTION VI

WETLAND RESTORATION: MAKING WETLANDS WORK AGAIN, 719

31 Introduction – Re-establishment of Wetland Functioning, 721 Edward Maltby 32 Restoration of Wetland Environments: Lessons and Successes, 729 Arnold G. van der Valk 33 Replumbing Wetlands – Managing Water for the Restoration of Bogs and Fens, 755 Russ P. Money, Bryan D. Wheeler, Andy J. Baird and A. Louise Heathwaite 34 Restoring Wetlands for Wildlife Habitat, 780 Dieter Ramseier, Frank Klötzli, Ursula Bollens and Jörg Pfadenhauer 35 Wetland Conditions and Requirements for Maintaining Economically Valuable Species: Waterfowl, Furbearers, Fish and Plants, 802 Lisette C.M. Ross and Henry R. Murkin SECTION VII SUSTAINABLE UTILISATION OF WETLANDS: BALANCING ECOSYSTEM FUNCTIONING AND HUMAN NEEDS, 819 36 Introduction – Sustainable Wetlands in a Global Context, 821 Tom Barker 37 Melaleuca Wetlands and Sustainable Development in the Mekong Delta, Vietnam, 829 R.J. Safford, Edward Maltby, Duong Van Ni and Nick P. Branch 38 Multiple Use of Wetlands in Eastern Africa, 850 Reint Jacob Bakema, Geoffrey W. Howard and Adrian P. Wood 39 Deterioration and Rehabilitation of the Lower Danube Wetlands System, 876 Angheluta Vadineanu

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Contents

40 The Pantanal of Mato Grosso: Linking Ecological Research, Actual Use and Management for Sustainable Development, 908 Wolfgang J. Junk, Carolina J. Da Silva, Karl Matthias Wantzen, Catia Nunes da Cunha and Flavia Nogueira 41 Wetlands for conservation and recreation use in the Norfolk and Suffolk Broads, 944 Tom Barker, Steve Crooks and Johan Schutten 42

Everglades and Agriculture in South Florida, 961 Robert H. Kadlec

43

Conclusions: Wetlands for the Future, 983 Edward Maltby and Tom Barker

Glossary, 1003 Index, 1007 Colour plates appear in between pages 530–531

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Preface

Wetlands are diverse ecosystems that link people, wildlife and environment in special and often interdependent ways through the essential life-support functions of water. Yet, once armed with the technology, human endeavour has focussed primarily on the large-scale dehydration of these landscapes, apart from exceptional and localised circumstances such as the creation of mediaeval fish ponds, more recent aquaculture developments, decoy habitats for hunting or aquatic gardens. Although considerably depleted in area compared with their historical extent, a new perspective of wetlands is now emerging and it is this change in attitude that underpins the philosophy, rationale and motivation behind the present text. The term, ‘wet land’ has been long used pejoratively, inferring land conditions that are less than ideal for the majority of practical purposes. ‘Wetland’ is a relatively new entry in dictionaries, even in the United States, where its more technical usage originated, probably in the 1950s in such publications as an inventory of wetland wildlife habitats by Shaw and Fredine in 1956. The term generally has been applied with an ecological, rather than any wider functional, connotation. Webster’s interpretation, for example, setting aside the plethora of recent technical definitions, states, ‘swamps and marshes, especially as an area preserved for wildlife’ (Merriam-Webster 2006). This restricted but common view of wetlands has supported a dichotomy between those areas that may, or even should, be altered for more directly ‘productive’ uses, and those that should or could reasonably become part of a network of protected sites. The basis of the latter designation is embedded in the more traditional thinking of nature conservation, emphasising species and communities (especially those that are rare, threatened or endangered) or exceptional examples of a particular ecosystem type. This rationale has supported

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some cutting edge scientific research on the fundamental ecology of individual species, communities and wetland types, together with improved understanding of the management requirements for maintenance of particular ecological characteristics, such as application of burning, grazing, water management, and the harvesting or control of wildlife populations. Conservation viewpoints, however, can overlook the much wider role of wetlands as parts of complex human and socio-economic landscapes, in which it is essential to consider ecology and economics together in a more coherent approach to decision-making, rather than as separate and in conflict. A generally held conservation-ecological perspective may view wetlands primarily as natural communities, with the management objective of maintaining the species, patterns and processes within individual wetlands. A more recent perspective is functional, viewing wetlands as ‘living machines that provide services to humans’ (Keddy 2000). This case has been argued for some time by Maltby (Maltby 1986, 1988; Maltby et al. 1994) because it puts wetland protection and management into the context of societal values such as water quality, flood risk reduction and fisheries support. Politicians as well as the general public can more easily evaluate these benefits against competing economic returns compared to the traditional nature conservation criteria. A strict interpretation of this view may infer that as long as particular functions are performed, the precise character of the ecosystem is of little significance compared with its utilitarian values. The two views, however, are not contradictory. Particular species and communities may have specific and even unique functional roles. They are, nevertheless, examples of different perspectives of wetlands. Such apparently divergent scientific positions on the significance of these ecosystems to society and our environment can contribute to

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x confusion and misunderstanding, especially in the implementation of appropriate policies. The paradigm presented in The Wetlands Handbook attempts to break down the artificial divisions between the natural science of wetlands and the societal criteria for their management. Such greater coherence is essential in deciding on their future; a future capable of harnessing their full, but often hidden and ignored, values. The editors thank Rosemary Maltby for tireless editorial assistance in the early stages of the project in managing contributors and reviewers. Vicky Cook manipulated manuscripts across different computer networks. Jos Verhoeven, Dennis Whigham and Mark Brinson, together with many unnamed reviewers gave their time freely to scrutinise manuscripts and provide advice. Delia Sandford’s patience and encouragement made it possible to complete the task of mobilising so many experts. Leendert Pons passed away before being able to see the final outcome of his labours. His enthusiasm for, and knowledge of, soils will be sadly missed.

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Preface R EFER EN CES Keddy P.A. 2000. Wetland Ecology. Principles and Conservation. Cambridge University Press, Cambridge, UK. Maltby E. 1986. Waterlogged Wealth. Earthscan, London. Maltby E. 1988. Wetland resources and future prospects – an international perspective. In: Zelazny J. and Feierabend J.S. (editors), Increasing Our Wetland Resources. National Wildlife Federation, Washington, DC, pp. 3–14. Maltby E., Hogan D.V., Immirzi C.P., Tellam J.H. and van der Peijl M.J. 1994. Building a new approach to the investigation and assessment of wetland ecosystem functioning. In: Mitsch W.J. (editor), Global Wetlands: Old World and New. Elsevier, Amsterdam, pp. 637–658. Merriam-Webster 2006. Merriam Webster’s Dictionary and Thesaurus. Merriam Webster Inc., Springfield MA. ISBN: 0877798516. Shaw S.P. and Fredine C.G. 1956. Wetlands of the United States. Their Extent and Their Value to Waterfowl and Other Wildlife. US Fish and Wildlife Service, Circular 39, 67 pp.

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Contributors

Michael C. Acreman Hydro-ecology and Wetlands, Centre for Ecology and Hydrology, Wallingford, Oxfordshire, UK Andy J. Baird Room 3.68, School of Geography, University of Leeds, Woodhouse Lane, Leeds LS2 9JT, UK Reint Jacob Bakema Freelance Rural Development Consultant, PO Box 5767, Kampala, Uganda Chris Baker Wetlands and Water Resources Management, Wetlands International Headquarters, Horapark 9, 6717 LZ Ede, The Netherlands Tom Barker Institute for Sustainable Water, Integrated Management and Ecosystem Research, Nicholson Building, University of Liverpool, Liverpool L69 3GP, UK Martin S.A. Blackwell Biogeochemistry of Soils and Water Group, North Wyke Research, Okehampton, Devon, EX20 2SB, UK Roland Bobbink B-Ware Research Centre, Radboud University, PO Box 9010, 6500 GL, Nijmegen, The Netherlands Ursula Bollens Landschaftsarchitekturbüro, asp Landschaftsarchitekten AG, Tobeleggweg 19, 8049 Zürich, Switzerland Samuel A.F. Bonnett Institute for Sustainable Water, Integrated Management and Ecosystem Research, Nicholson Building, University of Liverpool, Liverpool L69 3GP, UK Chris Bradley School of Geography, Earth and Environmental Sciences, University of Birmingham, Birmingham, UK

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Nick P. Branch School of Human and Environmental Sciences, University of Reading, Whiteknights, PO Box 227, Reading, RG6 6AB, UK Scott D. Bridgham Center for Ecology and Evolutionary Biology and Environmental Studies Program, 5289 University of Oregon, Eugene 97403, OR, USA Mark M. Brinson Biology Department, East Carolina University, Greenville, NC 27858, USA Roy Brouwer Department of Environmental Economics, Institute for Environmental Studies, VU University Amsterdam, The Netherlands Bernard Clément Unité Mixte de Recherches ‘Ecobio’ 6553, Centre National de la Recherche Scientifique, Université de Rennes 1, Campus de Beaulieu, 35042 Rennes Cedex, France Hadrian F. Cook School of Geography, Geology and the Environment, Kingston University, River House, 53-57 High Street, Kingston upon Thames, Surrey KT1 1LQ, UK Steve Crooks Phil Williams and Associates Ltd, 550 Kearny Street, 9th Floor, San Francisco, CA 94108-2404, USA Carolina J. Da Silva Mato Grosso State University, Cáceres Brazil Nancy B. Dise Department of Environmental & Geographical Sciences, Manchester Metropolitan University, John Dalton Building, Chester Street, Manchester M1 5GD, UK

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Contributors

Katherine C. Ewel USDA Forest Service, 60 Nowelo St., Hilo, H1 96720 USA; Present address: School of Forest Resources and Conservation, PO Box 110410, University of Florida, Gainesville, FL 32611 USA C. Max Finlayson Institute for Land, Water and Society, Charles Sturt University, PO Box 789, Albury, NSW 2640, Australia S. Georgiou Centre for Social and Economic Research on the Global Environment, University of East Anglia, Norwich, and University College London, London, UK Dave J. Gilvear School of Biological & Environmental Sciences, University of Stirling, Stirling FK9 4LA, UK Brij Gopal School of Environmental Sciences, Jawaharlal Nehru University, New Delhi 110067, India Ab P. Grootjans IVEM, Center for Energy and Environmental Studies, University of Groningen, Nijenborgh 4, 9747 AG Groningen, The Netherlands

Robert H. Kadlec University of Michigan, and Wetland Management Services, Chelsea, MI, USA Frank Klötzli ETH, Institute of Integrative Biology, Universitätstrasse 16, 8092 Zürich, Switzerland Gary A. Lamberti Department of Biological Sciences, University of Notre Dame, Notre Dame 46556, IN, USA Joseph S. Larson Environmental Institute, University of Massachusetts, Amherst, MA, USA Conor Linstead Institute for Sustainable Water, Integrated Management and Ecosystem Research, Nicholson Building, University of Liverpool, Liverpool L69 3GP, UK Edward Maltby Institute for Sustainable Water, Integrated Management and Ecosystem Research, Nicholson Building, University of Liverpool, Liverpool L69 3GP, UK Matthew P. McCartney International Water Management Institute, Addis Ababa, Ethiopia

Louise Heathwaite Centre for Sustainable Water Management, Lancaster Environment Centre, University of Lancaster, Lancaster LA1 4YQ, UK

Russ P. Money South East Region Water Resources Specialist, Natural England, Government Buildings, Coley Park, Reading RG1 6DT, UK

David V. Hogan Environmental Consultant, 291 Pinhoe Road, Exeter, Devon, EX4 8AD, UK

Henry R. Murkin Institute for Wetland and Waterfowl Research, Ducks Unlimited Canada, PO Box 1160, Stonewall, Manitoba R0C 2Z0, Canada

Geoffrey W. Howard IUCN East Africa Regional Office, PO Box 68200, Nairobi, Kenya Hans Joosten Institute of Botany and Landscape Ecology, University of Greifswald, Grimmer Strasse 88, D 17487 Greifswald, Germany Wolfgang J. Junk Working Group Tropical Ecology, Max-Planck-Institute for Limnology, 24306 Plön, p.b. 165, Germany

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Flavia Nogueira Federal University of Mato Grosso, 78070-030 Cuiabá, Av. Fernando Correa s/n, Brazil Catia Nunes da Cunha Federal University of Mato Grosso, 78070-030 Cuiabá, Av. Fernando Correa s/n, Brazil

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Contributors Jörg Pfadenhauer Vegetation Ecology, Technische Universitaet Muenchen, Am Hochanger 6, Building 219, 85350 Freising-Weihenstephan, Germany Gilles Pinay School of Geography, Earth and Environmental Sciences, University of Birmingham, Edgbaston, Birmingham B15 2TT, UK. Leendert J. Pons (deceased) Agricultural University, Wageningen, The Netherlands Michael C.F. Proctor School of Biosciences, University of Exeter, Geoffrey Pope Building, Stocker Road, Exeter EX4 4QD, UK Dieter Ramseier ETH, Institute of Integrative Biology, Universitätstrasse 16, 8092 Zürich, Switzerland K.R. Reddy Wetland Biogeochemistry Laboratory, Soil and Water Science Department, University of Florida, Gainesville, FL 32611, USA

Matthew Simpson WWT Wildfowl & Wetlands Trust, Gloucestershire GL2 7BT, UK

xiii Consulting, Slimbridge,

R. Daniel Smith Research Ecologist, Engineering Research and Development Center, Wetlands and Coastal Ecology Branch, 3909 Halls Ferry Road, Vicksburg, MS 39180, USA Julian R. Thompson Wetland Research Unit, Department of Geography, UCL, Pearson Building, Gower Street, London WC1E 6BT, UK Merritt R. Turetsky Department of Integrative Biology, University of Guelph, Guelph, Ontario N1G 2W1, Canada R. Kerry Turner Centre for Social and Economic Research on the Global Environment, University of East Anglia, Norwich, UK Angheluta Vadineanu Department of Systems Ecology and Sustainability, University of Bucharest, Splaiul Independentei 91-95, 050095, Bucharest, Romania

Curtis J. Richardson Duke University Wetland Center, Nicholas School of the Environment, Levine Science Center, Durham, NC 27708, USA

Panchabi Vaithiyanathan Divers Alert Network, 6, West Colony Place, Durhan, NC 27705, USA

Henri Roggeri IUCN National Committee of The Netherlands, Plantage Middenlaan 2-K, 1018 DD Amsterdam, The Netherlands

Isabel J.J. van den Wyngaert Alterra, PO Box 47, 6700 AA Wageningen, The Netherlands

Lisette C.M. Ross Institute for Wetland and Waterfowl Research, Ducks Unlimited Canada, PO Box 1160, Stonewall, Manitoba R0C 2Z0, Canada

Arnold G. van der Valk Department of Ecology, Evolution, and Organismal Biology, Iowa State University of Science and Technology, Ames, IA 50011-1020, USA

R.J. Safford BirdLife International, Wellbrook Court, Girton Road, Cambridge CB3 0NA, UK

Rudy Van Diggelen Ecosystem Management Research Group, Department of Biology, University of Antwerp, Universiteitsplein 1, B-2610 Wilrijk, Belgium

Johan Schutten Ecology, Environmental Science Directorate, Scottish Environment Protection Agency, Carseview, Castle Business Park, Stirling FK9 4SW, UK

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Duong Van Ni Hoa An Research Station, Can Tho University, Can Tho, Vietnam

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Contributors

Jos T.A. Verhoeven Landscape Ecology, Institute of Environmental Biology, Utrecht University, PO Box 80084, 3508 TB Utrecht, The Netherlands Melanie A. Vile College of Liberal Arts and Sciences, Department of Biology, Villanova University, 800 Lancaster Avenue, Villanova, PA 19085, USA Karl Matthias Wantzen Aquatic-Terrestrial Interaction Group, Institute of Limnology, University of Konstanz, Postfach M659, 78457 Konstanz, Germany Bryan D. Wheeler Department of Animal and Plant Sciences, University of Sheffield, Sheffield, UK

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Dennis F. Whigham Smithsonian Environmental Research Center, Box 28, Edgewater, MD 21037, USA John R. White Department of Oceanography and Coastal Sciences, Wetland and Aquatic Biogeochemistry Laboratory, Energy Coast & Env Building #3239, Louisiana State University, Baton Rouge, LA 70803, USA R. Kelman Wieder Department of Biology, 105 St. Augustine Center, Villanova University, 800 Lancaster Avenue, Villanova, PA 19085, USA Adrian P. Wood Wetland Action, 1070NB Amsterdam, The Netherlands

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Plate 1.7 The progressive contraction over time of Lake Chad owing to diversion of river water for irrigation and other uses (UNEP 2002). (Reproduced with permission from UNEP.)

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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(a) 5000

(b) 2000 1500

3000

Volume (106 m3)

Volume (106 m3)

4000

2000 1000 0 −1000

1000 500 0 −500

−2000

−1000

−3000

Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec (c) 4000

Volume (106 m3)

3000

2000

1000

0

River inflow

Rainfall

River outflow

Evapotranspiration

Groundwater recharge

Change in storage

Jan 87

Jan 86

Jan 85

Jan 84

Jan 83

Jan 82

Jan 81

Jan 80

Jan 79

Jan 78

Jan 77

Jan 76

Jan 75

Jan 74

Jan 73

Jan 72

Jan 71

Jan 70

Jan 69

Jan 68

Jan 67

Jan 66

Jan 65

−2000

Jan 64

−1000

Plate 6.7 The water balance of the Hadejia-Nguru Wetlands, northern Nigeria (January 1964–November 1997). (a) Mean annual inflows (+ve) and outflows (−ve), (b) Mean monthly inflows, outflows and change in surface water storage, (c) Monthly inflows, outflows and change in storage.

Plate 6.11 (a) Level–area relationship for a subcatchment of the Sussex Wildlife Trust’s Reserve on the Pevensey Levels, (b) Level–volume relationship for the same area, (c–f) Extent and depth (m) of inundation in the same area associated with different ditch water levels. (Source: Gasca-Tucker, in preparation.)

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Maltby-Plates.indd 3

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(a)

Sea

Fr

es

h

4m

Br

ac

kis

Sa

h

lt

(b) 1m

200

m

CaCO3 0.25%

3m

40 3H(TU)

(c)

Clay

200

m

Plate 8.2 Regional hydrological system of the Wadden Sea Island of Schiermonnikoog (a) and local hydrological system of freshwater dune slack (Kapenglop) influenced by through-flow of calcareous groundwater (changed after Grootjans et al. 1996a). (b) shows the depth of the decalcification front (CaCO3 50% >50% 93.6% 40% 66%

Africa Wetlands in Tunisia (84% in Medjerda catchment) (Hollis 1992) In parts of Tugela Basin; 58% in Mfolozi catchment, Natal.

15% 90%

Table 1.5 Half a century of wetland losses, catalogued by the US Fish and Wildlife Service. 1956 1983 1991 2000 2006

First report on wetland status and classification indicates substantial loss of wetland habitat for migratory waterfowl. First statistical wetlands status and trends report estimates rate of wetland loss mid 1950s to mid 1970s at 185 400 ha a−1 Second report. Update for period mid 1970s to mid 1980s estimates rate of loss declined to 117 400 ha a−1 Third Report covering 1986–1997 estimates rate of loss reduced further to 23 700 ha a−1 Fourth Report covering1998–2004 estimates gain overall in wetlands of 12 900 ha a−1 but still an overall loss of freshwater emergent marshes of 57 720 ha for the period.

specifically at the creation, enhancement and restoration of wetlands (Figure 1.10). Much of this has taken place on both active and inactive agricultural land (Dahl 2006). It is now estimated that, in 2004, there were 43.6 million ha of wetland in the conterminous United States, of which 95% were freshwater and 5% estuarine or marine.

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Shaw and Fredine 1956 Frayer et al. 1983 Dahl and Johnston 1991 Dahl 2000 Dahl 2006

Consequences of wetland ecosystem loss Most inventory work to date has focused on the recording and analysis of changes in the physical extent of wetlands. However, this does not reflect the important qualitative changes in wetland ecosystems that may result from both natural and human-induced impacts. The consequences

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19

The Changing Wetland Paradigm

Table 1.6 Changes in freshwater wetland area (thousands of ha) by type in the conterminous United States from the 1950s to 2004. Period

1950s

1970s

1980s

1998

2004

Forested Shrub Emergent vegetated Area of ponds

24 700 4 450 13 400 939

22 300 6 270 11 500 1 780

21 000 6 970 10 700 1 970

20 500 7 390 10 600 2 230

21 100 7 140 10 600 2 520

Hectares (millions)

Source: Dahl and Johnson 1991; Dahl 2000, 2006.

Area of US (conterminous) wetlands lost or gained 1998–2004 150 100 50 0 –50

r

te wa

p ee

nt

n

ba

Ur

D

me

l

ra Ru

lop ve de

ure

ult

c ilvi

S

ure

ult

ric Ag

r

he

Ot

Land use

Fig. 1.10 Area of wetlands lost or gained in the 48 conterminous states of the United States 1998–2004. Losses of wetlands are greater in industrialised areas, while mitigations are more frequent in rural districts. (Source: Dahl 2006.)

of wetland alteration are raising fundamental questions of irretrievable losses in human culture, other species, materials, information or economic benefit arising from ecosystem services (Costanza et al. 1997; Balmford et al. 2002). An example of the quantitative significance of wetland ecosystems is the potential value to the pharmaceutical industry and human wellbeing derived from indigenous knowledge of species of the peat-swamp forests of south-east Asia (Table 1.7). By analogy, the monetary value of any one species may range from millions to billions of dollars (Blum 1993). It would be misleading, however, to assume that dramatic losses of wetlands have passed without comment, even during the ‘heyday’ of agricultural development and new engineering techniques. In the English fenland ‘the profitseeking “adventurers” saw their drainage efforts resented and resisted by the ordinary Fenmen,

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whose common rights to grazing and fen produce were about to be expropriated’ (Bayliss-Smith and Golson 1992). The objections of wildfowlers, traditionally harvesting many thousands of ducks and geese for the London market, were particularly venomous (Darby 1983). Nonetheless, the sectoral economic interests of arable agriculture overwhelmed other considerations in the fens, just as they have done in many other parts of the world. Such overriding forces have continued to promote wetland degradation and loss without any properly informed debate of the issues. Some of the overriding factors are summarised in Table 1.8. The impetus for restoration Efforts to rehabilitate the River Rhine have been progressing for 25 years, following declines in water quality and habitat due to flow regulation

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20

EDWARD MALTBY

Table 1.7 Medicinal plants from the peat-swamp forests. (After Chai et al. 1989.) Family

Species

Plant form

Annonaceae Annonaceae Apocynaceae Araceae Davalliaceae

Fississtigma rigidum Mitrella kentii (Bl.) Miq. Alstonia spathulata Bl. Scindapsus perakensis Hook. F. Nephrolepis hirsutula (Forst.) Presl.

Woody climber Woody climber Tree Herb Fern

Medicinal value

Drink for treating fever For treating gonorrhoea (poultice from ash of stem) For shingles (leaves made into poultice) For easing pain caused by stings (leaf paste used) To stimulate lactation (drink made from boiled young shoots) Guttiferae Cratoxylum arborescens (Vahl) Bl. Tree For treatment of chicken pox (apply latex on rashes or skin disease) Leeaceae Leea sp. Shrub For scorpion and centipede bites, wasp and bee stings (apply paste made from young twigs and leaves to wound) Leguminosae Sindora leiocarpa Backer ex K. Hyne Tree Tonic (boil tap root and drink fruits boiled and mixed with other spices as a drink) Moraceae Ficus crassiramea Miq. Strangling fig For snake bite (apply paste on wounds – leaves, bark, roots) Myrtaceae Eugenia cerina Hend. Small tree Tonic (leaves) Myrtaceae Eugenia paradoxa Merr. Small tree For treating diarrhoea (leaves infused) Myrtaceae Eugenia zeylanica (L.) Wight Small tree Tonic Piperaceae Piper arborescens Roxb. Climber For treating rheumatism (plant boiled and drunk) Simaroubaceae Quassia spp Medium-sized trees To cure impotence and hypertension (boil tap root and drink)

Table 1.8 Examples of overriding factors promoting wetland loss. (After Dugan 1990 and Maltby 1991.) Science and information Imbalance of costs and benefits

Policy conflicts Institutional and management deficiencies

Ignorance of values of wetland functions. Products are ‘free’; costs of replacing them unseen. Need effort to measure non-market values and relate these to decision makers. Benefits accrue to developer; costs, in terms of harm to water, fisheries, or wildlife, to the wider community. The converse is also true: a farmer who restores a wetland may not see the benefits enjoyed by those downstream. Holistic planning with sympathetic intervention mechanisms are needed. Incentives from one governmental body often clash with those of another for example proposals for agricultural development might be in direct opposition with environmental policy. Insufficient attention given to wetland conservation issues, and a lack of funding, integration, management and experience among government agencies.

and embankment. River–floodplain transition zones, including marshes some distance from the active channel, were most affected (Buijse et al. 2002). Improvements were designed to increase species diversity in ecologically important reaches and to mitigate the toxic contamination of sediments, so that water could be suitable for a potable supply (Van Dijk et al. 1995). There have been marked improvements in water quality and, to a lesser extent, the sediments, which have led to increases in biodiversity. The rehabilitation

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of lost and degraded ecosystems, including the creation of new secondary channels, contributes to increased habitat diversity and heterogeneity (Nienhuis et al. 2002). The International Commission for the Protection of the Rhine, which oversees the project in conjunction with non-governmental organisations (NGOs), had projected that costs would amount to €5.5 billion up to 2005 (Buijse et al. 2002). Such restoration schemes are directed by assessment of the ecological state of the river–floodplain system,

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The Changing Wetland Paradigm including classification of the existing and potential hydrogeomorphological conditions, the position (height and width) of the floodplain relative to the river, and selected indicators of biodiversity (e.g. fish, aquatic macrophytes and invertebrates). Cooperation and a rigorous interdisciplinary approach have been identified as important factors influencing the success of the project (Buijse et al. 2002). The successful restoration of wetlands relies on a good understanding of their hydrology, and establishment of the conditions suitable for ecological succession and maturity of the system. These issues are complex, and need careful management if they are to be successful (see Chapters 32 and 33), but targets for conservation or ongoing harvesting of species for commercial reasons are important incentives for wetland restoration, and management must meet the particular requirements for provision of these services. Chapters 34 and 35 discuss the considerations and measures needed to ensure that targets are met. Increasingly, wetlands have featured in programmes and activities to raise standards of living and underpin the delivery of integrated and sustainable development (e.g. Pirot et al. 2000). The re-establishment of Melaleuca wetlands in the Mekong Delta, Vietnam (see Chapter 37) is just one example of many worldwide that demonstrate the capacity of often highly productive natural wetland systems to deliver multifunctional benefits of economic value (see also Chapters 25 and 38).

STRU C T U R E OF A N E W P AR ADIGM Contemporary issues in wetlands: an underlying rationale to improve understanding and management There are key underlying reasons for the recent rise of interest in the nature and future status of wetlands. Collectively they comprise the essential elements of a new paradigm for wetland science, based on a more holistic, interdisciplinary

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Fig. 1.11 Wetlands can be at the centre of a new, mature understanding of the priority concerns of society, linking natural and social sciences (expanded in main text).

approach than has been applied previously (see Figure 1.11). The cross-cutting perspective of the new paradigm places wetlands centrally in the implementation of the ecosystem approach (CBD 2005) that is, the integrated management of water, land and living resources to deliver sustainable development in an equitable way. As we move out from the focus on wetlands themselves, the new model highlights the linkages to be made between society and the natural environment on the one hand, and environmental management on the other, in order to achieve sustainable use of wetlands that is valuable for society as a whole. These linkages are made by way of the roles wetlands have in the water cycle, ecosystem functioning, spatial relationships and policies, and feed into management of water resources, use and conservation of wetland resources, connectivity and vulnerability in the landscape, social significance and the economic values of wetlands in providing ecosystem services. Application of this holistic approach can be made only by interdisciplinary collaboration between and within the natural and social sciences.

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Wetlands and the water cycle ‘Hydrology is probably the single most important determinant of the establishment and maintenance of specific types of wetlands and wetland processes’ (Mitsch and Gosselink 2000). Complex interrelationships with the hydrological cycle (Bullock and Acreman 2003) and essential dependency on water supply place wetlands centrally in some of the most contentious and urgent issues governing the appropriate management of water. Therefore, whilst wetland hydrology is a natural science, its relationship to the water cycle is highly relevant to societal concerns such as issues of water supply, resource allocation, water quality and flood risk. Wetlands have been cited variously as retaining water or increasing evaporation rates, reducing flood peaks, desynchronising the flood flow of tributaries, recharging or discharging groundwaters, and ameliorating low flows. Nonetheless, whilst wetlands are significant in altering the water cycle, the precise ways in which this can occur differ according to wetland type, position in the landscape, catchment characteristics and human interventions. The extensive database assembled by Bullock and Acreman (2003) indicates, for example, that whilst floodplain wetlands reduce or delay floods, some headwater wetlands can actually increase flood peaks. It is concluded that most wetlands evaporate more water than other land surfaces, and many reduce annual average river flows and decrease flow during dry periods. Nearly 40 years ago, the US Army Corps of Engineers (USACE), in its work along the Charles River, near Boston, Massachusetts, recognised the potential societal benefits of using natural wetlands for flood control. Contemporary proposals to alleviate flooding included the building of reservoirs, walls and dykes. An alternative was to simply protect the 3440 ha of wetlands as natural water storage areas (see Chapter 21). In 1972, a USACE study estimated that flood damage would increase by at least $3 million per year if 40% of the Charles River wetlands were destroyed, rising to $17 million annually if all the wetlands were

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lost (Horwitz 1978; Sather and Smith 1984). The retained wetland was valued at $1 203 000 per year, which was the difference between annual flood economic losses based on present land use and conditions, and projected flood losses by 1990. In 1983, USACE completed the acquisition and setting up of a protection regime for the basin wetlands of the Charles River (Weiskel et al. 2005). Whilst Acreman et al. (2003) report that floodplain wetlands on the River Cherwell, UK, reduced flood peaks downstream by more than 50%, there is now considerable evidence that this is a function that cannot necessarily be extrapolated to wetlands generally, and emphasises the need for objective ways of assessing, both individually and collectively, the probable functioning of wetland ecosystems (see chapters in Section 4). In particular, the multifunctional properties of most wetlands means that lack of performance in a single function should not necessarily limit the overall significance and value of the area. Water resources Wetlands are inevitably influenced by the growing competition for freshwater and the increasing human conflicts associated with its use (Gleick et al. 1994; Ward 2002). The amount of freshwater abstracted from natural systems for purposes other than the support of ecosystem functioning has probably increased some twenty times in the second half of the twentieth century (Gleick 1993; Abramovitz 1996; Postel et al. 1996) and the trend is largely unabated. Fundamental issues concerning the balance of water to support the direct needs of humans as opposed to natural ecosystems have been analysed in Falkenmark and Rockström (2004). The need to find an optimum balance between water utilised directly for human purposes (‘blue’ water) and that which delivers benefit through ecosystems (‘green’ and ‘indirect blue’ water) is embedded in a widened water resource context built on an integrated ecohydrological framework. This new perspective is based on rainfall, rather than runoff, as the basic freshwater resource. It recognises that both water

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The Changing Wetland Paradigm vapour and runoff support ecosystem functioning, that a multiscale approach from individual field or habitat to the river basin is essential, and that water quantity and quality dynamics need to be considered together (Falkenmark and Rockström 2004). Figure 1.12 illustrates one way of partitioning rainfall into ‘green’ and ‘blue’ flows, where rainfall supports freshwater functions in the four ecohydrological domains of: direct green use (ecosystem products); indirect green use (ecosystems); direct blue use (societal use e.g. industry, irrigation); and indirect blue use (ecosystem security). One estimate is that only 1% of global terrestrial rainfall flows though wetlands, but the yields of ecosystem goods and services are disproportionately higher than this (Balmford et al. 2002; Costanza et al. 1997; Falkenmark 1997; Falkenmark and Rockström 2004).

The implication of this is that wetlands multiply the value of global rainfall compared with other natural and man-made ecosystems. The importance of functioning An interaction of physical, chemical and biological processes and the particular elements of ecosystem structure result in ecosystem functions such as floodwater control and nutrient retention. Wetland processes have distinctive features associated with hydrology, biogeochemistry, productivity and food webs, which distinguish their behaviour from those operating in other ecosystem types. Differences in the performance of functions between one wetland and another result from variations in the rates of key processes, interactions of processes and ecosystem structure variability.

Output: evaporation

Total inputs (precipitation 100%) Grazing 18% Food 4% Wetlands (1%)

Other ecosystem services (incl. grasslands, forests, groundwater etc. 39%)

direct Catchment (Ecology Production Consumption Pollution)

indirect Seeping water

Flowing water

indirect

Aquatic ecosystems 5% Runoff 27%

direct Available for consumption and economic production (incl. industrial, domestic, food, tourism) 6%

Output: flow to sea

Fig. 1.12 Pattern of global blue and green water flows, showing direct human uses for green (supporting food production) and blue (supporting society) functions, and indirect uses, with green (ecosystem support) and blue (running or standing water) functions, and losses of the total to evaporation and sea. Wetlands, accounting for just 1% of global water outputs, provide a disproportionately high value of ecosystems services. Note that water flows cannot be unequivocally divided into functional types. (T. Barker; based on Falkenmark and Rockström 2004.)

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Ecosystem structure Geomorphology Hydrology Soils Fauna and flora

Processes Physical Chemical Bilogical

Wetland functions Hydrological functions Biogeochemical functions Ecological functions

Wetland dynamics level Societal benefit level

Services e.g. Flood control Water quality Food chain support

Ecological/environmental service webs

Life support

Market and non-market goods and social values

Goods e.g. Wood Plants Fish Birds

Economic products webs

Life support

Market and non-market goods and social values

Fig. 1.13 Physical, chemical and biological processes lie behind the provision of ecosystem services. (From Maltby et al. 1994.)

Wetlands are important for the provision of environmental and ecological goods and services, often called ecosystem services (MA 2005), that result from functioning (Figure 1.13). Wetland ecosystems generate services such as the supply and quality of water, controls on climate and pollution, and conservation of species and habitats. Goods include trees, birds and fish. Both services and goods provide direct and indirect benefits to human and wildlife populations and are important in environmental maintenance and quality. Aspects of water quality are dealt with in Chapters 17 and 20, and water quantity in Chapters 28 and 33, while Chapters 9, 10, 11 and 12 respectively deal with cycling and dynamics of nitrogen, phosphorus, carbon, and the

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interactions among physico-chemical variables. Chapter 15 discusses decomposition and mineralisation within wetlands. Wetland ecosystems possess attributes that include biological diversity and cultural heritage (see Chapters 13 and 14). All aspects of ecosystem condition, interactions and dynamics may be manifested as values (see Chapter 26). Whilst the functions performed by wetlands are of undoubted value to society, it is worth emphasising that they exist also in the absence of society, and are a normal part of the self-sustaining properties of an ecosystem (Brinson 1993). Recognition of the importance of wetland ecosystem functioning has caught the attention of the general public in developed and developing

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The Changing Wetland Paradigm countries through their experiences of the consequences of wetland loss, which may impinge directly on human welfare. This can take various forms, for example loss of fisheries, increases in flooding hazard, or declines in water quality. The dramatic decline in the sardine industry of the eastern Mediterranean was a direct but unanticipated consequence of the completion of the Aswan Dam, the physical changes it brought to the Nile Delta, and subsequent interactions with the marine ecosystem (Aleem 1972). In 1993 and 1995, Belgium, France, Germany and The Netherlands suffered widespread flooding in the catchments of the Rivers Rhine and Meuse. Economic losses were calculated at US$ 1 billion and US$ 3 billion respectively (Hausmann and Perils 1998). In 1997, a flood of the River Oder, which flows through Germany, Poland and the Czech Republic, was unprecedented in its depth and duration, killing over 90 people and flooding hundreds of villages (De Roo et al. 2003). In 2002, major floods occurred in Austria, the Czech Republic, Germany, Russia, Romania, Spain and Slovakia (Kundzewicz 2005), affecting major cities including Salzburg, Prague and Dresden. There were 100 fatalities, mostly in Russia, and losses were €15 billion (US$ 19 billion; Swiss Re 2003). The extent of recent flood damage in the United States (Sullivan and Galloway 2006), Europe (Preuss 2005) and Mumbai (Sturcke 2005) has been a stark reminder of the natural role of floodplains, and the hazards as well as high costs that may result from their development for alternative uses. In developed countries, it is increasingly difficult to obtain household flood insurance cover for residential developments on floodplains. This represents a change to the previously neglected relationships between human economies and ecosystem functioning. Throughout the developing world, however, extensive poverty means that direct human suffering can be the price paid for failure to recognise the importance of wetland functioning. Globally, human-induced changes in temperature and precipitation over the last 30 years are estimated to cause in excess of

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150 000 deaths annually (McMichael et al. 2004). Flooding is responsible for 40% of all natural disasters globally (French and Holt 1989), and the number of people at risk from coastal storm surges is predicted to rise from the present 75 million to 200 million by 2080, according to mid-range estimates (Stott et al. 2004; Blackwell and Maltby 2006). Conservation and management of biodiversity Wetland biodiversity is significant for at least three reasons: contributions to the global gene pool; support of species with special (particularly physiological and biochemical) adaptations to anaerobiosis, nutrient limitations, hydrological changes and other environmental stresses; and the presence of populations of some species of high economic and societal importance, especially in the case of communities in the developing world. Wetlands support very large numbers, and a rich diversity, of animal and plant species (Chapter 3). They are essential elements of the world’s major bird migratory routes and provide habitat for many endemic, rare and threatened species. Lake Balaton in western Hungary is a large shallow lake, the largest in central Europe. More than 70% of its fish species are endemic, it supports some 2000 species of algae, and over 1000 species of invertebrates. The lake provides a major wintering refuge for up to 70 000 geese, 30 000 ducks and 10 000 coots. At the same time it offers a commercial fish catch of 1200 tonnes per year. The complex of wetland ecosystems of the Danube Delta, covering some 5300 km2 also supports exceptionally rich bird populations. Among the 280 species that nest, rest or feed are tufted duck (Aythya fuligula), ferruginous duck (Aythya nyroca), red-crested pochard (Netta rufina), pygmy cormorant (Phalacrocorax pygmaeus), purple heron (Ardea purpurea), squacco heron (Ardeola ralloides), spoonbill (Platalea leucorodia), white pelican (Pelecanus onocrotalus), Dalmation pelican (Pelicanus crispus) and glossy ibis (Plegadis falcinellus).

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Mammals include otter (Lutra lutra), mink (Mustela lutreola), stoat (Mustela erminea), wild boar (Sus scrofa) and wild cat (Felis silvestris). Many of the species are either threatened in, or have been lost from, large areas of Europe. Despite extensive development of the Danube’s margins, some of the original floodplain ecosystems still survive. Reminders of the primeval landscape, floodplain forests such as those in the Gemenc area of Hungary provide habitats for birds that include the white-tailed eagle (Haliaeetus albicilla), black stork (Ciconia nigra), black kite (Milvus migrans) and night heron (Nycticorax nycticorax). All are indicators of the floodplain’s high ecological value, and are key targets in rehabilitation schemes for European rivers. Wetland species such as the otter, and many birds, are standard bearers of environmental quality, and indicators of societies’ ethical commitment to the species. Geneticists are interested in the potential of wetland plants and animals to improve cultivars and domesticated species; rice, oil palm and sago palm are just three wetland plants of enormous economic and social significance. The previously underestimated importance of paperbark (Melaleuca sp.) in the Mekong Delta is examined in Chapter 37. Iconic species and ‘unspoilt’ areas may generate revenue from tourism as well as maintaining aesthetic and ethical values (Chapters 40 and 41). Thus, tourism in the Northern Territory of Australia is linked in part to ‘the fascination with crocodiles, and based on attributes such as potential threat, danger, power, links with the prehistoric, and survivorship’ (Ryan 1998). It is economically significant that tourists prefer to see crocodiles within their natural ecosystems. Mbaiwa (2003) describes the Okavango Delta as one of Botswana’s leading tourist areas, mainly because of the rich wildlife resources and associated scenic character of the wetland. Wild harvests of lechwe and fish from the floodplains of East Africa, capybara and caiman in the Pantanal of South America (see Chapter 40), and the rich fisheries of the seasonally-flooded lands of the Mekong, are vital to the livelihoods of many local populations. Enthusiasm for ecotourism, however, should not mask the responsibilities

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associated with it. Tourism may also have important impacts on the ecosystems themselves. Unique human communities such as the Marsh Arabs (Madan) of Southern Iraq, the seasonally migratory Nuer and Dinka people of Sudan’s vast Sudd swamps, and pastoral tribes such as the Tuareg, Warbe´, Sonhabe´ and Peuhl of the inner Niger delta, are all examples of special cultural and socio-economic dependency on wetland environments and their biodiversity. External observers, who often argue their case on the grounds of sustainability, harmony and tradition, often view such relationships as ‘ideal’ and a priority for conservation. Such views, however, may ignore the aspirations of the inhabitants for improved provision of health, education and enhanced economic opportunities, which often are pursued through modification of the wetland environment, or movement away from it (see e.g. Chapter 38). The fundamental issue is how best to balance the needs of people, wildlife and environmental quality. Unquestionably, the conservation and ‘wise use’ of wetlands can play a significant part in meeting such diverse needs through the sustainable provision of goods and services. Decisions on where, and to what extent, wetlands may be altered to accommodate new balances should be based on sound scientific evidence and judgement. Biological reorganisation An issue of increasing significance is the spread of invasive alien species into wetlands, including plants such as mimosa (Mimosa pigra) in Vietnam, Australia and Africa, and Himalayan balsam or policemans helmet (Impatiens glandulifera) across Europe (e.g. Pysek and Prach 1995), and fish, of which some 160 species have been introduced to 120 countries (Balon and Bruton 1986). Prominent among these is the Nile perch (Lates niloticus) which, in Lake Victoria, has wiped out two-thirds of the indigenous small haplochromid fish species (Witte et al. 1992; Goldschmidt et al. 1993), and has adversely affected subsistence fisheries whilst supporting large-scale commercial fishing. Wetlands are particularly vulnerable to invasions, either because

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The Changing Wetland Paradigm of their former ecological isolation or the access to them facilitated by rivers. Spatial aspects Wetlands are naturally highly sensitive to disturbance, partly because of their particular spatial distribution. They are located in contact zones, interfaces or ecotones between land and water, which can expose them to maximum impacts from human activities and environmental change. The most common interfaces are at river and lake margins, and include more or less extensive floodplains. These are normally part of highly dynamic systems, resulting from flood pulses (see Chapter 40), seasonal fluctuations or other natural environmental cycles. They are often highly effective flood control systems (Acreman et al. 2000; Chapter 28) and buffers against land-based pollution. Many organisms, especially fish, rely on the natural flood regime to complete essential life cycle stages across these interfaces (Welcomme 1979; Moss 1998). Yet these are the zones most likely to be altered by flood control structures, or suffer impacts from alternative development pressures and the unmitigated effects of diffuse or point-source pollution. Wetland riparian buffer zones are examined in more detail in Chapter 19. The wetland complexes of deltas are a special case of spatial vulnerability. Often located at the bottom of very large catchments that may be shared by several countries, delta wetlands generally experience great cumulative effects from upstream, but may exist within national or other administrative boundaries that occupy only a small part of the drainage basin. Delta wetlands may be of disproportionate importance, such as in supporting fishery resources for other adjacent coastal or upstream nations. This is the case of the Danube, Rhine, Tigris-Euphrates, Mekong, Nile and other deltas of global significance. Connectivity issues The consequences of wetland degradation and loss may have direct and indirect effects on people, ecosystems and environments far distant from the point of impact. This may occur through linkages

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such as via migratory fish and birds, through the release of carbon dioxide, methane or nitrous oxide from drained peatlands, or through nutrient additions that change the ecology of water bodies downstream. The loss of gaseous carbon and nitrogen contributes to exacerbation of global warming, and hence has consequences such as sea level rise that has impacts on locations remote from the affected wetlands (see Chapter 11). The H5N1 virus, which causes avian influenza (bird flu), is currently prevalent among chicken farms in Asia, and has recently been found in Africa, probably following wild bird migrations (Ducatez et al. 2006). In the United States, fears of direct migration links with Asia have instigated a US$ 29 million programme to test wild birds for H5N1 in wetlands on the Bering Sea (Check 2006). There is doubt about the likelihood of transfer of the virus to humans by migrating birds, but healthy birds in China were found to be carrying the virus at the start of autumn migrations (Chen et al. 2006). Wetlands are an increasing part of transboundary agreements, for example the Danube Delta Biosphere Reserve (see Chapter 39) and the Mekong River Commission, international conventions such as Ramsar and the Convention on Biological Diversity, and transnational instruments (e.g. the EU Water Framework Directive). Policy framework Wetlands have their own dedicated international convention (Ramsar 1971), which has increasingly strong alliances with other global conventions, particularly the Convention on Biological Diversity (CBD), the United Nations Commission on Sustainable Development (CSD), and the United Nations Framework Convention on Climate Change, together with regional agreements such as, in Europe, the Birds Directive (79/409/EEC), the Habitats Directive (92/43/EEC) and the new Water Framework Directive (2000/60/EC). These, together with other wide ranging pieces of legislation, have resulted in an unprecedented level of potential political and legal force underpinning wetland conservation and management (see Table 1.9 for global and regional

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Table 1.9 Examples of global and regional governmental policy and treaties relevant to wetlands and wetland species. Major global policies and treaties Ramsar Convention UN Convention on Biological Diversity UN (Bonn) Convention on the Conservation of Migratory Species of Wild Animals (CMS) UNESCO Man and Biosphere Programme (MAB)

UNESCO World Heritage Convention BirdLife’s Global Important Bird Area Programme

Key site-based mechanism Ramsar Site designation, through Article 2.4 UN Framework Convention, which through Article 8 guides in situ conservation of species UNEP Framework Convention to enable such agreements as the African Eurasian Migratory Waterbird Agreement (AEWA) Biosphere Reserves, designated by individual states voluntarily often including protected areas designated under other international instruments World Heritage Site designation, some of which are important for biodiversity Inventory of the world’s most important bird areas. Not a legally binding instrument but often seen as a ‘shadow list’ for more formal designation of sites under national or international legislation

Major regional policies and treaties Africa Eurasia Migratory Waterbird Agreement (AEWA) EU Bird Directive Habitat Directive

Key site-based mechanism Flexible general flyway agreement under the CMS Special Protection Area (SPA)* system for birds described in Annexes Special Areas for Conservation (SAC)* for flora, fauna and habitats as described in Annexes Council of Europe Network of Protected Areas Includes the European Network of Biogenetic Reserves and the Pan-European Ecological Network implemented through national legislation Algeria Convention (Africa Convention on Conservation of Framework Convention, implemented through national legislation. Has Nature and Natural Resources) led to the establishment of many National Parks in African countries. No longer very active Bern Convention (Convention on the Conservation of Framework Convention, implemented through national legislation European Wildlife and Natural Habitats) leading to the Network of Areas of Special Conservation Interest (ASCI) (‘Emerald’ Network); modelled according to Natura 2000 criteria, but not legally binding Oslo and Paris Convention for the Protection of the Marine Not specifically aimed at site protection but has a special Protocol on Environment of the North-east Atlantic (OSPAR) North habitat protection and provisions to mitigate for changes in habitat American Wetlands Conservation Act character. Important instrument for USA, Canada and Mexico to protect, restore, or develop wetlands habitat; the act is not aimed at designating sites with a protected status Western Hemisphere Convention: Convention on Nature All American Convention for the protection of species and habitats, and Protection and Wildlife Preservation in the Western for requesting the designation of protected sites such as national Hemisphere parks, national reserves, wilderness reserves and national monuments. Not implemented Helsinki Convention on the Protection of the Marine Framework Convention recommends Baltic Sea Protected Areas Environment (HELCOM) Western Hemisphere Shorebird Reserve Network (WHSRN) Not legally binding, flexible site network for internationally important sites for migratory waders throughout North and South America Asia Pacific Migratory water bird Conservation Strategy Not legally binding, flexible site networks for various water bird groups Neo-tropical Migratory Bird Conservation Act Adopted by USA Government in 2000, focus on supporting the conservation of migratory birds and their habitat in the Western Hemisphere. No specific site designation arrangements *The SPA and SAC networks together form the Natura 2000 network, a legally binding protected areas network for EU states (including those in the process of accession). Non-compliance may lead to formal legal cases brought to the European Court. Source: From Boere and Taylor (2004).

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The Changing Wetland Paradigm examples, and also Chapters 21 and 24 for discussion of the formalisation of wetland protection in the United States). There is increasing pressure on governments to protect and ensure the ‘wise use’ of wetlands within their sovereign territory. Powerful NGOs and smaller action groups are increasingly effective in reminding Contracting Parties of their national as well as international obligations through lobbying of the Conferences of Parties and pressure on the funding sources enabling ‘unwise use’. For example, Day et al. (2006) have provided independent technical commentary on relevant proposals by the Spanish National Hydrological Plan to transfer up to 1050 km3 of water per year from the lower Ebro River to sites in the north of the country and the southern Mediterranean coast. They concluded that the plan was unsound for a number of reasons, which included threats to important habitats and species protected under existing European legislation, the reduction in water and sediment flow, which put at risk the sustainability of the ecosystems and human communities of the delta (a Ramsar site and one of the most important wetland areas in the western Mediterranean), the need to counteract ongoing subsidence and sea-level rise, and the widespread public opposition from affected stakeholders. The planned water transfers were denied European funding and have now been abandoned by the Spanish government, at least for the time being, and replaced by alternative proposals, including desalination, water re-use and improved demand management. Targeted policy Some countries, including the United States, Canada and Uganda have specific and dedicated wetland policies. In 2004, the United States introduced an initiative that established a federal policy beyond the principle of ‘no net loss’ of wetlands. The new policy aims to increase both the quality and quantity of wetlands, and to restore, improve and protect more than 1.2 million ha in 5 years. After the loss of over half the US wetlands

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since colonial times (Tiner 1984), this marks a major turning point. The most significant factor explaining the turnaround from loss to gain has been due to the creation of artificial freshwater ponds. Pond area increased by 282 000 ha from 1998 to 2004, whilst natural freshwater wetlands continued to decline. Although forested wetland area increased by 223 000 ha, this was due mainly to the maturation of small trees previously not classified as forest (Dahl 2006). The reversal of wetland losses may be only the illusion of actual recovery of the resource. This is because there is still no assessment of the quality, functioning or condition of the stock, or of the changes that might have taken place in monitoring criteria. In particular, the replacement of naturally functioning wetlands by created wetlands may result not only in the reduction of existing functions (e.g. habitat for particular species or food chain support) but also a change in the balance of functioning, for example, resulting from the greater proportion of open water systems compared with wetlands dominated by dense emergent vegetation. The latter are likely to play a much more effective role, for example, in buffering diffuse pollution, owing to the structure of the ecosystem, and because of their location in the landscape (see Chapter 19). The policy of the Canadian Federal Government with respect to wetland conservation has recognised the importance of an approach that highlights their wider functional importance. As early as 1991, the stated objective was to: ‘promote the conservation of Canada’s wetlands to sustain their ecological and socio-economic functions, now and in the future’ (Government of Canada 1991). The ecological and socio-economic functions recognised are listed in Table 1.10. The clear challenge in the implementation of this policy is to achieve the appropriate balance of functions, especially where conflicts occur. Most obvious are those relating to peat harvesting and agriculture, which potentially are at odds with most, if not all, the ecological functions.

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Table 1.10 Functional rationale cited in Canadian and Ugandan wetlands policy. Canada Ecological Functions Water recharge Natural shoreline protection from wave action and erosion Natural flood reduction and control Important role in oxygen, evapotranspiration and climatic cycles Waterfowl, flora, furbearers, reptiles and fish habitat Refugia for rare and endangered species Biodiversity preservation Natural carbon store Natural sink for pollutants Nutrient source for connected waters Soil and water conservation Socio-Economic Functions Resource for hunting, trapping and fishing Tourism and recreation; Major domestic peat energy source Domestic source of peat for horticulture and agriculture Harvesting of hardwood and pulpwood Agriculture Aesthetic resource Scientific research resource Natural heritage Uganda Ecological functions Maintenance of the water table Prevention of erosion Reduction in extremes of flow Sediment traps Wildlife habitats and centres of biological diversity Socio-economic functions Plant products Fishing Cattle grazing Water supply Nutrient and toxin retention Tourism

Canadian wetland policy sets clear goals • Maintenance of the functions and values derived from wetlands throughout Canada; • no net loss of wetland functions on all federal lands and waters; • enhancement and rehabilitation of wetlands in areas where the continuing loss or degradation of

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wetlands or their functions have reached critical levels; • recognition of wetland functions in resource planning, management and economic decisionmaking with regard to all federal programmes, policies and activities; • protection of wetlands of significance to Canadians; • recognition of sound, sustainable management practices in sectors such as forestry and agriculture that make a positive contribution to wetland conservation while also achieving wise use of wetland resources; and • utilisation of wetlands in a manner that enhances prospects for their sustained and productive use by future generations. The goals are compatible with the ‘wise use’ requirements of the Ramsar convention but were also influenced by the broader thinking of the World Conservation Strategy (IUCN 1980), which identified wetlands as key life support systems, in tandem with agricultural lands and forests. They fit also within the ecosystem approach of the more recent Convention on Biological Diversity (CBD 2005) and the recommendations of the Millennium Ecosystem Assessment (MA 2005). Uganda Uganda’s national wetland policy is the first in Africa. Wetlands are recognised in the National Constitution as ecosystems of great value to the people. President Museveni, who had spent considerable time in the Ugandan swamps when in opposition to the previous regime, issued an edict in 1986 that there should be no further drainage of wetlands until a national policy was developed. The National Wetlands policy sets five goals: optimum use of wetland resources, termination of practices that reduce wetland productivity, maintenance of biological diversity, maintenance of wetland functions and values, and integration of wetland concerns into the planning and decision-making of other sectors. There is clear recognition of the need to integrate wetland management with all other aspects of development

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The Changing Wetland Paradigm strategies and activities. It is emphasised that wetland conservation can only be achieved through ‘a co-ordinated and co-operative approach involving all the concerned people and organisations in the country, including the local communities’. The text of the Ugandan policy (Republic of Uganda 1995) identifies ecological and socioeconomic functions of wetlands, similar to the Canadian policy (Table 1.10). The wide distribution of wetlands means that a large proportion of the population has access to their use, although their value is not always appreciated. Conversion to single sector use results in the loss of wider benefits from traditional uses, more restricted ownership of the resource, reduced economic flexibility, increased risk from crop pests and reduced human health. Decision-making regarding permitted or prohibited activity rests with District Authorities, who are also empowered to exercise ‘reasonable discretion’ on the exact area of wetlands to be developed within the context of the wetlands policy and ‘other prevailing policies on natural resources and the environment’. This encapsulates a developing world position in which flexibility and compromises may be essential elements of meeting societal priorities. European Water Framework Directive The Water Framework Directive (WFD: Directive 2000/60/EC on establishing a framework for community action in the field of water policy) is the most substantial piece of European water legislation to date. It requires all inland and coastal waters to reach ‘good status’ by 2015. It involves a river basin approach and the setting of stringent environmental, especially ecological, targets for surface waters. Wetlands are not a primary emphasis within the Water Framework Directive, although their water needs are addressed clearly where they are dependent on surface or groundwaters. There is a dual approach to wetlands. On the one hand, the intention is to protect the integrity of important (protected area) wetlands, but also it is recognised by the European Commission, NGOs, and the agencies responsible for implementation

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that wetlands overall will be important in delivering the aims of the WFD (e.g. Defra http://www. defra.gov.uk/environment/water/wfd/). The European Parliament and Council agreed on the WFD in 2000. It promises to have a major impact on water management in Europe, and some sections are particularly relevant to wetland management, notably: Article 1: Establishment of a wetland protection framework. Article 4a: An emphasis on ecological quality as well as water quality. Article 5: The characterisation of wetlands in the context of their river basins. Article 8: The assessment of waters moving in and out of wetlands in terms of flow rates, chemical quality and ecological potential. Article 13: The development of river basin management plans. The emphasis on river basin management and the assessment of ecological quality will encourage more complete assessments to be made of wetland functions, and the benefits they provide, in the context of an entire catchment. Wetlands can contribute effectively to the delivery of the new European water policy embodied in the Water Framework Directive, but, despite a comprehensive guidance document, there is still limited evidence that European member states are developing formal worked strategies for wetlands in relation to the Directive. Some key objectives of the WFD may be being undermined by political expediency and reluctance to take the steps necessary to ensure fundamental improvements in water quality (Moss 2008). The Horizontal Guidance Document (EU 2003) suggests that EU member states establish a Programme of Measures to facilitate wetland creation, restoration and management projects to improve the provision of ecosystem services and help achieve the Directive’s overall goals. Wetland economics Costanza et al. (1997) emphasised the disproportionate contribution of wetlands to the natural capital provided by the world’s ecosystems

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cropland grass/rangelands forest temperate/boreal forest tropical lakes/rivers freshwater wetlands swamps/floodplains tidal marsh/mangroves shelf coral reefs seagrass/algae beds estuaries open ocean

Gas regulation Disturbance regulation Water supply Waste treatment Habitat/refugia

0

0

5000

10 000

15 000

Food production

20 000

1994 US$ ha–1 a–1 Fig. 1.14 Annual global value (US$ ha−1) of ecosystem services indicate the disproportionate importance of wetlands. (Data from Costanza et al. 1997.)

(Figure 1.14). Balmford et al. (2002) reviewed the costs and benefits of conversion of wild habitats to other uses, for example woodland into logs, and mangrove into aquaculture. They concluded that conversion was always harmful in overall economic terms. The authors’ conservative estimates indicated that investment in conservation and management of natural systems, rather than their conversion, is likely to pay back to the economy at least 100 times the investment, and be always much more cost-effective than investment in conversion. Conversions of natural systems into industrial production were motivated by the accrual of profits, but these benefits were received by the developer rather than the local or wider community, as had been the case prior to conversion (Balmford et al. 2002; see also Chapter 26). There is still much to be done before such comprehensive values of natural ecosystems are included in new decisionmaking practice. Balmford et al. (2002) also cite the analysis of van Vuuren and Roy (1993) who reported that, for freshwater marshes in Canada, the total economic value was considerably higher when the wetlands remained intact rather than converting them to agriculture ($8800 versus $3700 ha−1). The difference was attributed partly to the subsidised cost of drainage and partly to the wide social benefits of wetlands, including hunting, trapping and angling.

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Environmental economists are devoting considerable efforts to quantifying the relative economic and social values of wetlands (see Soderquist and Lindahl 2003; Turner et al. 2003). The development of robust and reliable economic arguments is essential for the improvement of decisions that affect wetlands. It remains necessary to link effectively and unambiguously scientific knowledge of how wetlands work, with the economic advantages they bring to both developed and developing countries. This topic remains a major objective for future collaborative research. Wetlands may be the most cost-effective means of delivery of a wide range of human, wildlife and environmental benefits. Based on these and other related issues, four interrelated topic areas are highlighted as examples.

Conservation and management of wetland ecosystems as a means of maintaining wide-ranging multifunctional benefits to society Many river floodplains are being altered and lost as a result of alternative water management. In Africa, this is a particular problem associated with large-scale irrigation schemes that take no account of the downstream impacts on wetlands or the benefits they provide to various sections

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The Changing Wetland Paradigm of society. Barbier and Thompson (1998) simulated various scenarios for major irrigation schemes in the Hadejia-Jama River basin in Nigeria. They concluded that the benefits from normal floodplain inundation exceeded those gained from irrigated agriculture. This means that where river regulation has already occurred, it is necessary to introduce flood releases to minimise the loss of benefits resulting from reduced downstream flooding. In other cases it was advised to avoid further expansion of largescale irrigation. In the developing world, development of intensified irrigation agriculture on normally dry land has been one of the greatest causes of wetland loss, since the waters that naturally feed wetlands are diverted or impounded. The comparison of the present value of net economic benefits in Nigeria of the Kano River irrigated agriculture project and the Hadejia-Nguru wetlands, from which water flow would be derived, is a clear illustration of

the scale of ‘opportunity cost’ associated with the shift from multifunctional floodplain ecosystems to single economic sectors (Table 1.11). The difference is particularly stark when expressed in terms of the volume of water used, but even this comparison may underestimate the scale of difference because other benefits such as groundwater recharge to the Chad formation aquifer, which serves numerous villages in the surrounding region, were not included in the analysis (Barbier et al. 1997). It is important to realise that whilst some of the benefits attributed to wetlands may be replaceable, this may be achieved only at greater overall cost to the community or wider society and, in some cases, in a very different form. Whilst there are alternative ways of achieving the improvement of water quality attributable to wetlands, and of delivering flood protection, there is no other way of maintaining the particular biodiversity directly dependent on the

Table 1.11 Comparison of value of net economic benefits, Kano River project and Hadjiea-Nguru wetlands, Nigeria under different financing options*. Discount rate

8%

8%

12%

12%

Period

50 years

30 years

50 years

30 years

278 100 4450

256 300 4100

190 000 3000

184 000 2930

381 233

351 214

260 158

253 153

109 000 0.3

101 000 0.3

74 500 201

72 400 195

14.5 0.00004

13.4 0.00004

9.94 0.0268

9.65 0.0260

1. Total (N’000)† H-N wetlands Kano River Project 2. Per hectare (N/ha)‡ H-N wetlands Kano River Project 3. Per water use (N/m3)§ H-N wetlands Kano River Project 4. Per water use (US$/1000 m3)§ H-N wetlands Kano River Project *7.5 Naira = 1 US$ (1998/1999 prices).

†Based on a total net benefit from agriculture, fuelwood and fish production attributed to Hadejia-Nguru wetlands and total net benefits of irrigated crop production from the Kano River Project. ‡Based on 230 000 ha cropland, 400 000 ha forest and 100 000 ha fishing, and total production area of 730 000 ha for Hadejia-Nguru wetlands and a total crop cultivated area of 19 017 ha in 1985/6 for the Kano River Project. §Assumes an annual average flow into the wetlands of 2549 Mm3, annual water use of 15×103 m3 per ha for the Kano River Project.

Source: Adapted from Barbier et al. (1993).

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wetland environment. Similarly, the archaeological and scientific (environmental and palaeoenvironmental) information stored in wetland stratigraphy is unique. Once information such as pollen sequences, biochemical markers and organic macro remains is changed or destroyed it cannot be recreated elsewhere. The evidence is often highly site-specific, especially for preserved archaeological remains. Additional examples are given in Chapters 37, 38, 39 and 42.

objectives of the European Water Framework Directive can be found in the Wetlands Horizontal Guidance Document developed as part of the Common Implementation Strategy of the WFD (http://www.wfduk.org/about_wfd/3-3/view). Floodplain characteristics have major controls on the flooding process (Table 1.12) and measures can be applied to achieve flood risk reduction using the natural floodplain system (Table 1.13). Further examples are explored in Chapters 6 and 19).

Wetland management as a mechanism for achieving environmental benefits at least cost, such as sustainable development through flood control, groundwater recharge, and the control of diffuse or point-source pollution

Restoration and creation of wetlands as an instrument for recovery of environmental quality, biodiversity and improvement in human welfare

Recent research has addressed all of these issues. A pertinent example is the publication by the European Commission of the Ecoflood Guidelines: How to Use Floodplains for Flood Risk Reduction (Blackwell and Maltby 2006), which reflects increasing concerns over flood risk, and the imminent Floods directive (Directive of the European Parliament and of the Council on the Assessment and Management of Floods; SEC(2006) 66) currently under discussion in the European Parliament. Further guidance on the use of wetlands to improve water quality and other

There is growing interest worldwide in the active restoration and creation of wetlands to realise significant ecological and societal benefits. Chapters in this volume by Safford et al. (Chapter 37) and Kadlec (Chapter 20) illustrate this for both the developing world (Mekong Delta, Vietnam) and the developed world (United States). However, a particularly topical example is illustrated in the case of the restoration of the marshlands of southern Iraq (see Maltby, Chapter 31). Within the last two decades alone, some 20 000 km2 of former marshland had been converted to an arid landscape, resulting in the loss of livelihood of

Table 1.12 Floodplain functions and processes (Blackwell and Maltby 2006). Functions

Processes

Hydrological functions (water quality related) Floodwater regulation River base-flow maintenance Sediment retention

Floodwater storage; increase in river discharge capacity Groundwater discharge Sediment deposition and storage

Biogeochemical functions (water quality related) Nutrient retention

Carbon accumulation

Plant uptake of nutrients; Storage of nutrients in soil organic matter; Adsorption processes in soil; Precipitation; Retention of particulates. Gaseous export of N (denitrification and ammonia volatilisation); Vegetation harvesting; Soil erosion. Accumulation of organic matter and formation of peat

Ecological functions (habitat related) Ecosystem maintenance Food web support

Provision of diverse habitats; Provision of habitat microsites. Biomass production; Biomass import; Biomass export.

Nutrient export

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The Changing Wetland Paradigm

35

Table 1.13 Natural flood risk reduction measures (Blackwell and Maltby 2006). Measure

Qualitative description of the measure

Protection of existing naturally functioning river and floodplain systems Flood bypasses

The existing storage capacity of the river system is maintained and valuable ecosystems are protected New river bypasses, including new floodplains with wetland or floodplain ecosystems. Also called green rivers Enlarges the effective river floodplain Enlarges the storage capacity of a floodplain and leads to enlargement and restoration prospects for a floodplain Increases the storage capacity of a floodplain

Removal/lowering of minor embankments Setting back of embankments (Re)construction of stagnant water bodies such as isolated channels and oxbows in the (former) floodplain Development of manageable flood detention polders, which should preferably be used as extensive grassland or floodplain forest Floodplain excavations Changes in land use in the catchment area (e.g. reforestation) Restoration of floodplain vegetation (Re)construction of meanders (Re)construction of flowing side channels Re-meandering the river course or allowing spontaneous river morphological development Removal of flow restrictions

Increases the storage capacity of a floodplain

Enlarges the effective river floodplain Promotes retention of water in a catchment area Increases the storage time of water on a floodplain Increases the storage capacity of a river channel and decreases a river’s slope Increases the storage capacity of a channel area and increases the water conveyance capacity through a river section Increases the storage capacity of a river channel

Alleviates unwanted flooding in some areas and purposefully relocates this to designated areas. Increases river flows downstream with managed storage areas used for habitat creation Rejuvenating or removing vegetation with a high hydraulic Only ecologically beneficial if the management of the vegetation roughness supports the development of a stable and viable ecosystem Removal or lowering of groynes or other hydraulic obstacles in Allows more dynamics in river level fluctuations, decreases a river/ the river channel valley roughness coefficient

several hundred thousand marsh dwellers, the marsh Arabs (Maltby 1994; Partow 2001). The reduction in discharge and magnitude of the flood pulse of both the Tigris and Euphrates resulting from dam construction in Turkey, as well as control structures in Iraq, enabled the Saddam regime to more easily complete drainage works in the marshes. This resulted not only in misery, control and displacement of entire human communities, but also the loss of important wetland habitat, ecosystem functions and probable extinction of some species. An aspect of marshland loss that is yet to receive significant attention is the linkage between the marshes and the marine environment, and the consequences first of the disruption, and second of the re-establishment of

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the hydrological and ecological continuum from the river basin to the open sea (Al-Yamani et al. 2007; Maltby 2006). Regime change in Iraq has provided an opportunity to restore at least part of the original marshland for the benefit of people, wildlife and environmental quality. There has been rapid recovery of wetland habitat as a result of both unregulated as well as officially sanctioned reflooding (Richardson et al. 2005; Richardson and Hussain 2006) but it is still not clear what will be the extent of marshland that can be restored, based on the competition for limited water resources. In 2005, reflooding occurred in over half the area of former marshland present in the 1970s. Although significant numbers of former marsh dwellers have returned

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EDWARD MALTBY

0.4

Reconstructed Actual data (1902–1995) 2σ error bars Reconstructed (50 Year lowpass)

0.0

–0.4

–0.8 1400

Fig. 1.15 Marsh Arabs have returned to the reflooded marshes at Suq Al-Shuyukh. Image shows reeds harvested for fodder. (Photo E. Maltby.)

to the reflooded landscape (Figure 1.15), many are located at the edge of the wetland. Relatively few occupy islands and traditional reed structures (mudhifs), once such a characteristic feature of the marshlands. Wetland restoration has certainly re-established options for the recovery of wildlife and ecosystems, and has also made possible the recovery of traditional culture, including the supply of fish, and the basis of support for livestock and raising of crops. Nevertheless, it is clear from interviews carried out by Maltby (unpublished) in 2004 that the Madan are not wedded to the hardships of life in the marshlands, no matter how sustainable and idealistic this might look from the outside. They are anxious to benefit from improved economic welfare, transport, education and medical facilities. This will mean inevitable changes to the original wetland landscape. Wetland restoration can contribute to enhanced human welfare in Mesopotamia, but this will be only part of an integrated solution to the management of land, water and living resources in the catchment (see Chapter 31). Dealing with climate change On a global scale, the importance of wetlands in releasing, sequestering and storing carbon (see Chapters 11, 12 and 18), in providing water

Maltby-C001.indd 36

1500

1600

1700 Year

1800

1900

Fig. 1.16 The original ‘hockey stick’ graph (Mann et al. 1998), showing temperature (°C) on the Y-axis. The zero line shows the calculated deviation from a calibration mean for the years 1902–1980.

(Chapters 7, 8 and 17), and in protecting the human environment from flooding (see Chapters 6 and 28) have come into sharper focus in recent years because of their complex roles in both the amplification and the mitigation of the effects of anthropogenic climate change. Correlations between the burning of fossil fuels, the increasing concentrations of CO2 in the atmosphere, and the rise in global mean temperature via the ‘greenhouse effect’ are accepted by consensus of the overwhelming majority of climate scientists. The famous ‘hockey stick’ graph of Mann et al. (1998; see Figure 1.16) was criticised by others (McKitrick 2005; McIntyre and McKitrick 2005) but, after further studies (Bürger and Cubasch 2005; Rutherford et al. 2005; Thejll and Schmith 2005), has now become generally accepted (Moberg et al. 2005; D’Arrigo et al. 2006), and even extended to 10 000 years BP (IPCC 2007). Political decisions about measures that can mitigate the potential changes, which may be severe and far-reaching, are progressing, but meaningful change is slow and increasingly hampered by short-term vested interests (Monbiot 2007). The role of wetlands in future environmental policy should not be underestimated. Whilst there is debate over whether

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The Changing Wetland Paradigm the world’s peatlands in their present condition will have positive or negative effects on global warming via greenhouse gas exchange, there is no doubt that their loss would rapidly release significant amounts of CO2 into the atmosphere, previously accumulated over millennia. Alternatively, the rejuvenation of peat formation on degraded surfaces, or initiation of new cycles of peat development as a result of hydrological regeneration, may result in rates of CO2 capture that exceed any counter-effect of methane evolution (Maltby and Immirzi 1993). There is less certainty regarding the balance of radiatively-active gases associated with other wetlands where, for example, the production by denitrification of nitrous oxide (approximately 300 times more potent than CO2) may exceed any benefit of carbon sequestration. Nevertheless, there may be major water quality benefits associated with the removal of nitrates, and wider biodiversity and amenity values associated with such functioning wetlands. There is considerable need for a more integrated scientific approach to examination of the roles of wetlands in mitigating or amplifying the processes and effects of climate change. Wetlands are at the cutting edge of the challenges facing interdisciplinary science and ecosystem management, where in particular we need to overcome the boundaries of the sectoral and often too narrowly focused disciplines. They provide some of the most pressing tests of society’s ability to resolve the fundamental conflicts among the many diverse interest groups of the world’s human community. R E F E R E N CE S Abramovitz J.N. 1996. Imperiled Waters, Impoverished Future: The Decline of Freshwater Ecosystems. Worldwatch Paper 128, Worldwatch Institute, Washington, 80 pp. Acreman M.C., Barbier E.B., Birley M., Campbell K., Farquharson F.A.K., Hodgson N. Lazenby J., McCartney M.P., Morton J., Smith D., et al. 2000. Managed Flood Releases – Issues and Guidance. Report to DFID and the World Commission on Dams. Centre for Ecology and Hydrology, Wallingford, UK.

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2 Global Distribution, Diversity and Human Alterations of Wetland Resources DE NNI S F. WH I G H AM Smithsonian Environmental Research Center, Edgewater, USA

IN T R O D U CT ION This chapter outlines the distribution and diversity of wetland resources and examines the types of human activities that have resulted in widespread alteration. The term ‘wetland’ has a variety of meanings, and examples are given of the terminology used to describe different types of wetlands. The processes responsible for the formation and persistence of wetlands are indicated and consideration is given to human perceptions of these ecosystems and threats to them from both natural and anthropogenic activities.

D E FI NI T IO N S AN D CL AS S IF ICATION Many definitions of wetlands have been developed (Maltby 1991; Dugan 1993; Mitsch and Gosselink 2000; Tiner 1996) and some of them have been changed over time, particularly in the US, in response to an increased understanding of wetland ecology and to political arguments (National Research Council 1995). Two examples of wetland definition demonstrate the range of considerations that have been used to describe their characteristics. The most widely known and internationally adopted wetland definition was developed for purposes of providing international protection for waterfowl across the widest possible range of wetlands. The Ramsar Convention on The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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Wetlands of International Importance Especially as Waterfowl Habitat, published in 1971, gives: … wetlands are areas of marsh, fen, peatland or water, whether natural or artificial, permanent or temporary, with water that is static or flowing, fresh, brackish or salt, including areas of marine water the depth of which at low tide does not exceed six metres.

Another widely used wetland definition was developed by the US Fish and Wildlife Service for purposes of bringing consistency to an ongoing national debate related to inventory, regulation and conservation of wetlands in the United States (Cowardin et al. 1979; Mitsch and Gosselink 2000; Tiner 1996). The Cowardin et al. definition, developed as part of a classification system, is: … lands transitional between terrestrial and aquatic systems where the water table is usually at or near the surface or the land is covered by shallow water. For the purposes of this classification wetlands must have one or more of the following three attributes: (1) at least periodically, the land supports predominantly hydrophytes; (2) the substrate is predominantly undrained hydric soil; and (3) the substrate is nonsoil and is saturated with water or covered by shallow water at some time during the growing season of each year. (‘nonsoil’ simply means ‘not soil’ e.g. rock, undecomposed litter, or the sediments of a water body too deep for the growth of rooted plants (usually >2 m–3 m)).

Both definitions communicate the importance of vegetation and water in the identification, development and persistence of wetlands, and the Cowardin et al. definition also recognises the importance of substrate conditions. The two descriptions collectively identify the essential elements of a robust definition of wetlands: hydrology,

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vegetation and substrates. Most wetland definitions are not, however, particularly useful when there is a legal requirement to determine whether or not a given habitat should be classified as a wetland or when wetland delineation is required. The issues of wetland identification and delineation have been particularly contentious in the United States and many efforts have been directed toward providing a definition that can be used to develop methods to identify and delineate wetlands (Tiner 2000). Tiner emphasised that a robust wetland definition that can provide effective guidance toward wetland identification and delineation must include the three elements listed above (hydrology, vegetation, soil). He also suggested that a tiered approach be applied in application of identification procedures. Obvious wetlands can be identified by rapid assessments using vegetation, hydrology or soils alone, while wetlands that are more difficult to identify require additional effort and multiple indicators of these elements. No matter how wetlands are defined, it is common practice to use widely accepted terms to describe different types. Dugan described globally distributed wetland types based primarily on geomorphic position (Dugan 1993). Estuaries, mangroves and tidal flats are wetlands and wetland habitats associated with coastal features that are tidally influenced. Floodplains and deltas are systems that contain wetlands associated with rivers with various flooding regimes. Marshes, lakes, peatlands and forested swamps are terms used by Dugan to describe wetlands associated with non-tidal inland habitats. Maltby (1991) also described general wetland types based on broad features such as dominant vegetation (marshes, swamps), soil characteristics (peatlands), geomorphic features (floodplain wetlands, lakes, estuaries and lagoons), geographic location (mangroves, Nipa swamps (Nipa palm, Nipa fructicans), and tidal freshwater swamp forests) and human activities (artificial wetlands). Mitsch and Gosselink (2000) recognised seven types of wetlands based on whether they were associated with coastal (tidal) or inland (non-tidal) habitats. Coastal wetlands included tidal salt marshes, tidal freshwater

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marshes and mangroves. Inland wetlands were freshwater marshes, northern peatlands, southern deepwater swamps and riparian wetlands. While the wetland terms used by Dugan (1993), Maltby (1991) and Mitsch and Gosselink (2000) are widely recognised, the reader should be aware that the use of common terms can lead to confusion because names for a single type of wetland can vary from one language to another (Scott and Jones 1995). At times, common terms used to describe types of wetlands can be confusing. The Ramsar definition given above, lists fens and peatlands, suggesting that they are separate types of wetlands. Gore (1983) defined fens as peat accumulating wetlands that receive rainwater and drainage from surrounding mineral soil and usually support marsh-like vegetation. He defined peatland as a generic term for any wetland that accumulates partially decayed plant matter (Gore 1983), indicating that fens are a type of peatland. It is preferred to describe all active peat-forming wetlands as mires, and separate them into fens and bogs according to whether they are fed predominantly by groundwater or rainfall. Despite the danger of confusion it is nevertheless useful to examine the salient characteristics of the most common categories found in the literature. Marshes Marshes (Figure 2.1) are wetlands dominated by herbaceous vascular plants, the stems of which emerge above the water surface. Marshes occur in areas that are frequently or continuously inundated with water and they are most often associated with mineral soils that do not accumulate peat. Typically, dominant plant species in marshes are reeds, rushes, grasses, and sedges that are characterised by thin ‘grass-like’ leaves. Marshes, however, can also contain a wide variety of plant species with many different life forms. Freshwater tidal marshes, for example, can be dominated by annuals and perennials that range in leaf form from grass-like to broadleaved (Simpson et al. 1983). Marshes have been studied extensively because of their importance as waterfowl habitat (Weller 1994) and many

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Distribution and Diversity of Wetland Resources

Fig. 2.1 The author standing in a stream channel at low tide in a freshwater tidal emergent wetland along the Delaware River (USA). Tidal amplitude is approximately 3 m. The dominant emergent species in the foreground is Nuphar advena (Yellow waterlily). (Photo by Robert Simpson.)

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Fig. 2.2 Richard Hauer in a freshwater swamp dominated by Acer rubrum (Red Maple) and Nyssa aquatica (Water Tupelo) on the Pearl River in Mississippi (USA). (Photo by the author.)

different types of marsh have been recognised both in the US (e.g. prairie potholes, playas, salinas, salt marshes, brackish marshes, freshwater tidal marshes, vernal pools and Carolina bays) and elsewhere (Semeniuk and Semeniuk 1995). Peatland (often called mire) This is a generic term for any wetland that has at some point accumulated partially decayed plant matter because of incomplete decomposition, usually to a depth less than 30 cm (Figure 2.2). The term ‘mire’ refers to those peatlands in which peat formation is still active. Many terms have been developed to describe peat-forming wetlands, particularly in Europe (Money). Fens can be dominated by herbaceous or woody plant species. Bogs are peatlands dominated by herbaceous or woody species, but they differ from fens because the water chemistry resembles that of precipitation and the peat is usually formed by the slow decomposition of mosses, especially species of Sphagnum. A complex and diverse terminology (Heathwaite and Göttlich 1990; Verhoeven 1992; Glooschenko et al. 1993) has been developed in different countries to describe different types of bogs.

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Fig. 2.3 Seiichi Nohara in the Akaiyachi Mire (bog), central Honshu, Japan. Sphagnum dominated peat has accumulated to a depth of 150 cm. Below the Sphagnum peat are layers of Moliniopsis and Phragmites dominated peat overlying a layer of volcanic ash and sand with wood and Phragmites remains. Four Sphagnum species dominate the existing peat mat. Abundant vascular plants are Phragmites australis, Ilex crenata var. paludosa, Vaccinium oxycoccus, Moliniopsis japonica, Rhynchospora alba and Sasa palmata. (Photo by the author.)

Swamps Wetlands that are covered intermittently or permanently with water, and are dominated by trees or shrubs, are swamps (Figure 2.3). In essence,

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DENNIS F. WHIGHAM

swamps are marshes that are dominated by woody vegetation. Swamps, like marshes, are diverse and occur in many habitats from the temperate zones to the tropics (Lugo et al. 1990). Swamps that are associated with rivers typically have inorganic substrates (Sharitz and Mitsch 1993; Junk 1997). Other swamps develop in areas with little or no connection to flowing streams and rivers, and may develop peat substrates (Richardson 1981; Richardson and Gibbons 1993). Swamps represent a particularly threatened type of wetland in many parts of the world because of exploitation for wood products and conversion to agricultural lands. There have also been attempts to classify wetlands based mostly on hydrologic and hydrogeomorphic characteristics, thus eliminating some of the confusion associated with the use of common names based mostly on vegetation characteristics. Scott (1989) and Scott and Carbonell (1986) developed a widely cited classification system (Table 2.1), for neotropical wetlands. They used dominant vegetation type to describe three wetland types (08, 18, 19) but location within the landscape, or a combination of landscape position and vegetation, were used to classify most wetlands. Cowardin et al. (1979) developed a system to classify wetlands in the US that is hierarchically based on geomorphic features (marine, estuarine, riverine, lacustrine, palustrine) hydrologic conditions (subtidal, intertidal, tidal, perennial streams, intermittent streams, limnetic, littoral) and several modifiers, such as substrate conditions and dominant vegetation type (Figure 2.4).

W E T L AN D DIS T R IB UT IO N Wetlands occur over wide range of altitudes and latitudes, from the tropics to the arctic, and from below sea level to alpine environments (Figure 2.5). They occur in almost all habitats, from tidally influenced coastal landscapes that flood once or twice daily, to endorheic basins in interior areas of continents that most often have little or no standing water but are transformed into productive wetland systems when rainfall occurs. The

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Table 2.1 Classification of wetlands used in the Directory of Neotropical Wetlands (Scott and Carbonell 1986). Identification of wetland types is based on geomorphic features (e.g. estuaries), habitat type (e.g. freshwater pond), vegetation type (e.g. mangrove), and substrate conditions (e.g. peat bogs). Vegetation type refers to the numbering system used by Scott and Carbonell (1986). Vegetation Description type 01 02 03 04 05 06 07 08 09 10 11 12 13 14 15 16 17 18 19

Shallow sea bays and straits Estuaries, deltas Small offshore islands, islets Rocky sea coasts, sea cliffs Sea beaches (sand, pebbles) Intertidal mudflats, sandflats Coastal brackish and saline lagoons and marshes, salt pans Mangrove swamps, brackish forest Slow-flowing rivers, streams (lower perennial) Fast-flowing rivers, streams (upper perennial) Riverine lakes (including oxbows), riverine marshes Freshwater lakes and associated marshes (lacustrine) Freshwater ponds (1.60 1.42–1.60 1.25–1.42 1.08–1.25 0.90–1.08 0.30 0.25–0.30 0.20–0.25 0.15–0.20 0.10–0.15 30

30

200

3000 g m−2 a−1). This level of production is common in wetlands. Table 18.3 shows net primary productivity (NPP) for a variety of natural wetlands (Aselmann and Crutzen 1989). When compared with estimated values of NPP for other ecosystems (Table 18.4; Melillo et al. 1993), it can be seen that NPP of natural wetlands can be ranked as high as that of temperate and tropical forests. The total biomass stored in natural wetlands is lower globally owing to the total area occupied by natural wetlands relative to other ecosystems. The productivity of reed marshes and mangroves in back-swamps declines rapidly as the vegetation of the fringes removes some of the nutrients from the floodwater, which is also progressively diluted with rainwater. This is the background to the processes of bio-filtering wastewater and the use of wetlands as buffer zones, but means also that the growth of mangroves and reed marshes may be significantly enhanced by small ditches, allowing the floodwater to enter the back-swamps directly. Higher tidal ranges yield more nutrients, which are clearly demonstrated by increases in Spartina alterniflora production on salt marshes along the USA Atlantic coast. In summarising production losses due to the presence of fringing vegetation, Wassink (1975) estimated a fall in production rates from 30 t ha−1 a−1 to 10 t ha−1 a−1 in temperate areas and,

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Table 18.3 Net primary productivity (NPP) from natural wetlands. Climate change region latitudes are: polar: 65–90°; boreal: 55–65°; temperate: 30–55°; tropical: 0–30°. Abbreviations: dm = dry matter, Mha = million hectares and Tg = 1 terragram = 1 million tonnes (Aselmann and Crutzen 1989). Wetland type

Climate region

Bogs

Polar Boreal Temperate Tropical

21 104 42 20

Total

187

Fens

Swamps

Marshes

Floodplains

Polar Boreal Temperate Tropical

Area (Mha)

54 62 32 –

Total

148

Polar Boreal Temperate Tropical

– 1 10 102

Total

113

Polar Boreal Temperate Tropical

– – 17 10

Total

27

Polar Boreal Temperate Tropical

– – 8 74

Total

82

Lakes

12 Total

569

at lower temperatures and in arid regions, rates below 10 t ha−1 a−1. In tundras and deserts a fall to below 2 t ha−1 a−1 was estimated. Nitrogen, phosphorus and carbon in alluvial soils and floodplains Figure 18.5 shows the likely fate of C, N and phosphorus (P) in the environment of a managed floodplain (excluding primary wetlands). Two hydrological scenarios are illustrated: one for mean stage of the river, the other at ‘mean annual flood’. In an agricultural context, the sensitivity

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NPP range in dm (g m2 a−1) 100–300 300–700 400–800 600–1200

NPP in dm (Tg a−1) 20–60 310–760 170–340 120–240 620–1400

100–300 400–700 400–1200 –

50–160 250–430 130–380 – 430–970

– 500–1000 700–1500 1500–3000

– 0–10 70–150 1530–3060 1600–3220

– – 800–2000 1500–4000

– – 140–340 150–400 290–740

– – 800–1800 1500–2500

– – 60–140 1110–1850 1170–1990

400–800

50–100 4160–8420

of nutrient behaviour is conditioned particularly by grazing, but also by supposed ‘misuses’ of floodplain agricultural land for arable agriculture. It is also important in management systems to consider the kind of flooding and its probable frequency. For example, historic water meadows represent a controlled diversion of water across the floodplain, permitting trapping of the silt and sand fraction of suspended sediment, the possible decrease in P that is mobile in the water, and the potential, although hitherto unproven, diminution of nitrogen water through particulate trapping and plant uptake (Cook et al. 2004).

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399

Wetland and Floodplain Soils Table 18.4 Model estimates of annual NPP for potential terrestrial vegetation. Values were converted from g C m−2 a−1 by multiplying by 2.22 (i.e. assuming dry matter contains 45% C) (Melillo et al. 1993). Ecosystem type

Climate region

Tundra

Polar desert/alpine Wet/moist Boreal Boreal Temperate coniferous Temperate mixed Temperate deciduous Temp. broadleaf evergreen Tropical deciduous Tropical evergreen Arid Mediterranean Short Tall Temperate Tropical Total

Woodland Forest

Shrubland Grassland Savanna

Area (MHa)

NPP range in dm* (g m2 a)

NPP in dm* (Tg a)

500 470 630 1 220 240 510 350 320 460 1 740 1 450 140 470 360 680 1 370 11 560

0–480 75–940 200–930 275–960 460–1560 510–2370 180–2170 715–2220 720–3100 900–3160 13–1000 70–1400 160–970 300–1700 150–1740 200–1745

0–2 400 350–4 420 1 260–5 860 3 350–11 710 1 100–3 740 2 600–12 090 630–7 600 2 290–7 100 3 310–14 260 15 660–54 980 190–14 500 100–1 960 750–4 560 1 080–6 120 1 020–11 830 2 740–23 900 36 440–184 040

*Dry matter.

Topographic Floodplain feature Land Arable use Inundation Average

Inputs of N and P 2

Drainage ditch

Floodplain meadow

Floodplain ‘Basin’

Grazing Average

Infrequent

Common

WT

L Accumulation of C, N, P

N, P, and C from flooding

L

Common

Rare

Mineralisation of N & P

Mineralisation of N & P 2 WT

1 Mineralisation/ Immobilisation of N & P

Flood embankment

Grazing

Removal of C from Denitrification Redistribution of crop N, P, C through Grazing 2

WT

Nature levèe

River channel

L

WT

Accumulation of C, N, P in topsoil

L

Key: 1 Riverstage at mean flow 2 Riverstage at mean annual flood (Q2.33) L Significant leaching occurs WT Watertable at stage 1

Fig. 18.5 Schematic of nitrogen, phosphorus and carbon behaviour under different floodplain management systems. (Cook 2007.)

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Deeper flows across the floodplain can still deposit and hence enrich the soil with significant amounts of sediment. With this are associated nutrients and carbon. Equally, there remains the risk of mobilising sediment, particularly from banks where arable agriculture has replaced primary wetland grazing. Conversely, the minimal restoration of arable with grass should, over decades, sequester carbon in topsoil against global warming. Subsoil processes, including denitrification in shallow water tables, are vaunted for the reduction of nitrate, and hence this process is promoted in the use of buffer strips of perhaps only a few metres along watercourses in reclaimed wetlands (White et al. 1998; Blackwell et al. Chapter 19). There is no known comparable process for attenuation of P. Although the latter is complexed in soils and taken up by vegetation, it cannot readily be lost to the atmosphere, and is mobile in solution beneath intensive arable plots (see Richardson and Vaithiyanathan, Chapter 10). There is the potential for soils and geological materials to complex with P, and hence buffer groundwater flowing towards wetlands and other water bodies. Phosphorus-free buffer strips are claimed to be beneficial in preventing P loading of water bodies (http://www.ag.ndsu.edu/pubs/ h2oqual/watnut/nm1298w.htm). Table 18.1 demonstrates that the form and function of wetland soils relies upon hydrological variables, and infers that water table height is a major factor that has to be manipulated in order to achieve management goals. The unsaturated zone above a water table is a soil horizon that experiences oxygen diffusion along the larger cracks and pores that also enable root penetration. Typical desirable depths of freeboard drainage (this is the vertical distance between topographic surface and free water surface observed in an adjacent ditch) for secondary (grazing) marsh are 0.3–0.5 m. Intensive grassland becomes possible at around 0.5 m or deeper freeboard drainage. Freeboard in alluvial marshlands in eastern England is typically maintained through a network of ditches with densities typically in the order of 6.5–8.5 km km−2 (Cook and Moorby 1993),

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although on the climatically wetter Tadham Moor in Somerset, the post-drainage and enclosure ditch network is 15 km km−2, twice the density found in eastern England. Deep underdrainage can increase the freeboard to between 0.8 and 1 m during the growing season (Cook 1999), permitting arable production unless there is serious flooding. Where reed beds are concerned, it is desirable to flood the soil to a depth of around 0.5 m for the cultivation of reeds and saw sedge. Methanogenesis Microbial repiration produces CO2 which, together with low-molecular acids, acts as a terminal electron acceptor at very low oxidation–reduction potentials, generating CH4 (swamp or marsh gas). This may dissolve in water or be released as a gas, especially at high temperatures (e.g. in the tropics or during warm seasons) and in freshwater ecosystems. There is concern over the conditions leading to the release of CH4 because it is an important greenhouse gas, but the process is natural, and important in peatlands, as the cause of floating rich-fens (Van Wirdum 1982) and the floating peat islands or ‘plaurs’ (Rudescu et al. 1964) typical of the Danube Delta. These create nesting places for birds, for example pelicans in otherwise very wet marshes, and thus support biodiversity. Gleying and reduction Gleying arises from redoximorphic processes, its features are the colour patterns in a soil due to loss (depletion) or gain (concentration) of pigment when compared with the soil’s matrix colour. The defining process is the reduction of iron from ferric Fe3+ to ferrous Fe2+ forms, or the reduction of manganese from black Mn3+ or Mn4+ to manganous (and soluble) Mn2+, or the reverse (oxidations) of these processes, which may be coupled with their removal, translocation or accrual, resulting in a soil matrix colour controlled by the presence of Fe2+ (NSSC 2002). Reduced wetland soils have grey to bluish or greenish-grey colours, due to dissolved Fe2+ ions, in contrast to oxidised soils, which contain

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Wetland and Floodplain Soils prominent brown or reddish-brown coloured ferric oxides. Upper layers may show thin (0.5–1.5 cm), brown, oxidised O2-diffusion layers at the soil–floodwater interface, due to oxidation of Fe2+ into Fe3+ ions. The aerenchyma tissue of many wetland plants permits O2 transport into the roots, and creates narrow, oxic, brownishcoloured areas around them in the otherwise grey (reduced) matrix. In reclaimed soils, roots of agricultural crops produce a similar effect, and soil mottling occurs where there is a mixture of reduced (greyish) and oxidised (orange or yellow) iron species. Leached ferrous ions from elevated upland soils may move through reduced subsoils to more oxidised soils in valleys (Schlichting and Schwertmann 1973). Here they may accumulate as ‘iron rocks’ and ‘bog iron’ deposits, which historically have been used as building materials and industrial ore. As redox potentials decline and electronaccepting substances are reduced (Table 18.2), new, reduced electron-losing substances are formed. These, respectively, with redox values, are: H2O (water, at +600 to +400 mV), N2O, N2 and NH4+ (N components, at +250 mV), Mn2+ (manganous, at +225 mV), Fe2+ (ferrous, at +120 mV), S2− (sulphide, at −75 to −150 mV) and CH4 (methane, at −250 to −350 mV) Acid soils and pyrite formation Some peat deposits and more generally organic surface layers are associated with particular coastal and alluvial deposits known as potential acid sulphate (PAS) materials, which pose very specific management problems when drained (Maltby et al. 1996). Oxidation of PAS materials leads to the development of acid sulphate soils (ASS), which release sulphuric acid and produce extreme acidity, not only in the soil but also in drainage water. There are estimated to be 12 million ha of PAS and ASS soils worldwide, mostly in the tropics. Acid sulphate soils have a ‘sulphuric horizon’ within a depth of 50 cm that shows acidity from oxidation of sulphides. The layer is at least 20 cm thick, is of pH of less than

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401

3.5, and has jarosite mottles not surrounded by brown mottles or, if peaty, has more than 0.5% sulphate, with or without jarosite (Pons 1989b). In many places in the world, pyrite (FeS2) muds have been or are being drained for rice cultivation and other crops. The only totally reliable means of preventing acidification and release of toxic mineral substances in this type of wetland soil is to maintain a high water level, preventing oxidation of the sulphidic layers (Satawathananont et al. 1991; Maltby et al. 1996). Acidity can be neutralised by application of lime but this is generally too expensive for most agricultural activities, and may be impossible in remote tropical floodplain or coastal regions. Alternatively, certain trees in the genus Melaleuca grow well under extremely acid conditions. Re-establishment of Melaleuca forest is possibly the only viable option for the productive use of acid sulphate soil areas with or without peat cover. Plantation of Melaleuca or other wetland types provides a means of preventing the further oxidation of PAS materials and the development of severe acidification in areas such as the Mekong Delta, Vietnam (see Safford et al. Chapter 37). Waterlogged soils, infiltrated by saline and brackish water containing SO42− ions and iron substances will form and accumulate pyrite according to Equation 18.4 (Van Breemen 1976; Pons and Van Breemen 1982): 1 Fe2O3 + 4SO42− + 8CH2O + O2 → 2 2FeS2 + 8HCO3− + 2H2O

(18.4)

The six conditions necessary for pyrite formation are: • the presence of sulphur (derived from sulphate-ions); • the presence of iron (derived from mineral deposits); • an anaerobic soil; • the presence of easily decomposable organic matter in the soil; • sufficient permeability to allow SO42− ions to move through the soil; • a periodic inundation of floodwater containing some oxygen.

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These conditions are met in some saline and brackish, waterlogged soils, often in estuarine and deltaic environments. The composition of the various iron minerals that form in the process of pyrite oxidation is largely determined by the redox potential and the pH of the soil material. Soils on well drained levees will rarely form pyrite because restricted flooding increases the rate of aeration. Nevertheless, wetland soils in backswamps below mean water level and with abundant brackish water will accumulate amounts of pyrite that are potentially dangerous because they may result in acid sulphate soils if they become oxidised. The conditions are common, too, in marshes dominated by reed and rush for example Phragmites australis and Scirpus maritima (Dent 1986). The most important prerequisite for high pyrite concentrations in the mud is, however, the speed of accretion of the waterlogged soils. When a soil liable to pyrite formation is accreting rapidly, it means that large quantities of non-pyritic deposits will accumulate simultaneously, so that the total pyrite in these soils will stay low. On high pyrite producing soils at low rates of sediment supply, a mud containing a greater proportion of pyrite will result. Pyrite muds occur in slowly silting mangrove, mature mangrove and reed marsh wetland soil-profiles, usually in back-swamps where the amounts of pyrite may attain concentrations of 15% by mass or 100 kg m−3 (Dent and Pons 1995). Accumulation of pyrite takes place under neutral pH conditions. Mangrove and reed marsh soils with pyrite, if they are not aerated or drained, have pH-values varying between pH 5.5 and pH 7. These soils produce fertile sites for hydrophytic vegetation because they receive abundant nutrients, do not experience limiting conditions for growth, and are never short of water. Thus they may support very high net dry matter production and can contribute significantly to food-web support for human communities. They may be suitable places for wildlife reserves. During the geogenetic phase (Stage A, above), sediments are deposited with a particular content of primary carbonates (calcium [Ca2+] and magnesium [Mg2+]). If CaCO3 is present in the waterlogged

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soils, calcium bicarbonate (Ca(HCO3)2) is formed, which is soluble and easily leached from the soil (Van der Sluys 1970); the amount of water and the concentration of dissolved CO2 control the process. Decomposing organic matter, present in the A horizons, produces CO2 which, together with atmospheric CO2, is dissolved in the floodwater. Consequently, the water filtering through the soil may contain comparatively high concentrations of CO2 and HCO3− ions. The concentrations of CaCO3 ultimately found in the wetland soil developed on sediments will be a result of the balance between lime inputs from flood deposits and the loss of CaCO3 by leaching. This lack of residual CaCO3 will exacerbate any problems arising from pyrite and formation of acidsulphate muds. Oxidation and acid sulphate soils Acid sulphate soils are well known in England, the Netherlands and indeed worldwide. In the tropics, restoration of mangrove vegetation is very difficult because mangrove plants are not able to establish on severely acid soil. It was mainly the acidification of the soils in the Mekong delta, south-east Asia, that prevented the recovery of the mangrove and kajeput (Melaleuca) vegetation. Stands of Melaleuca leucadendron were lost through defoliant application during the Vietnam war (see Safford et al., Chapter 37). After drainage, anaerobic muds are soil-forming material in which the potential acidity exceeds the acid-neutralising capacity. The potential acidity of PAS soils is caused by pyrite that will oxidise after drainage as air penetrates into a totally reduced sulphidic mud, producing acidity and jarosite (Equation 18.5): FeS2 +

15 1 5 O2 + K + + H 2O → 4 3 2

4 1 SO24 + + KFe 3 (SO4 )2 (OH)6 + 3H + 3 3 FeS2 +

5 H2O → FeO.OH + 2H2SO4 2

(18.5)

(18.6)

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403

Wetland and Floodplain Soils At pH-values 1.5 mm) fibre particles decreases. Peats are classified into three categories: fibrists (less than one third decomposed) hemists (one to two-thirds decomposed) and saprists (more than two thirds decomposed) (Mitsch and Gosselink 2000), the remainder being identifiable plant fibres. Fibric peats (e.g. Sphagnum-rich) are relatively young and are light in colour in comparison to the other peat types. Fibric peats have a relatively high water-retention capacity and, in general, are quite acidic due to the lower degree of decomposition. Hemic peats are older and more decomposed than fibric peats, and originate generally from reeds, sedges, and other non-moss type plants. Sapric, or humic peats, are the oldest and most decomposed peats, with colours ranging from dark brown to black. Normally, sapric peats were the first peats formed in the filling of basins, and are the most dense and colloidal, thus they take in less water but retain it more strongly than other peats (Cantrell 1991). Peat formation may occur by below-ground accumulation of organic matter in originally mineral wetland soils and by accumulation of litter on the soil surface. With increasing, longlasting waterlogging, plant communities may add so much underground organic matter into the soil through their roots that a mineral soil will transform into peat. Most eutrophic peats (e.g. Phragmites peats) belong to this group, especially the peats formed under brackish conditions. The freshwater ‘wood peats’ of the Dutch lowland peats are also formed in this way (Pons 1992).

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Wetland and Floodplain Soils

407

In wetland soils of humid climates, under poor and declining drainage conditions and decreasing free water inflow (back-swamps and freshwater marshes), root decomposition in the soils decreases and A horizons (the top layer of soil below the litter layer) of wetland soils are replaced by peaty topsoils. Increasing amounts of organic matter are accumulated and the formation of either eutrophic or mesotrophic peat starts, depending on nutrient supply, but when rain water is the only available source of nutrients and the climate is sufficiently wet, oligotrophic peat domes eventually develop (Moore and Bellamy 1974). Raised bogs are peat deposits that fill entire basins, and form where rainfall is between 700 mm a−1 and 1000 mm a−1. They are raised above groundwater levels and therefore receive any inputs of nutrients from precipitation. Blanket bogs form where rainfall is greater than 1250 mm a−1 and, along with low temperatures, allow the peat to literally ‘blanket’ very large areas. Figure 18.6 shows a deep peat soil profile.

S O IL S IN M AN AGE D W E T L AN D AR E AS Hydric soils may remain within the definition despite changes to their physical form that might result from reclamation or construction of irrigation systems, or the imposition of irrigation or deep drainage. The soils of southern England provide useful examples. The Somerset Moors, UK (Figure 18.7), have been subject to moderate land drainage. Much of the area remains as ‘secondary’ wetland, despite attempts in recent decades to deep drain and reclaim it for more intensive agriculture, including arable (Cook 1998). Historical management had led to improved regional drainage through the period 1770–1840, especially the creation of new rectangular fields with ditches. In areas where fields mostly result from drainage of mineral alluvium, boundaries may be curvilinear, reflecting original drainage channels, including meandering

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Fig. 18.6 Exposed peat profile at Moorhouse National Nature Reserve, UK. (Photo by S. Bonnett.)

rivers across the primary marsh. Straight ditches were dug in order to increase the drainage densities, and these are also most typical of reclaimed peatlands (Cook and Moorby 1993; Cook 1994; Rippon 1999). Figure 18.7 (Findlay et al. 1984) shows a range of soils derived from marine and river alluvium and fen peats, due to marine and river environmental incursions in eu-fibrous peat soils in grass sedge peat, and the even more humified Adventurer’s series. Turbury Moor association soils are the remnants of raised bogs, and the name reflects the ancient right to dig peat and extraction, which continued until modern times. These peats, formed by ombrogenous bogs, are non-calcareous and occur below 6 m ordnance

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HADRIAN F. COOK ET AL.

High level drain

6m 814c 851a

3m

Pumping station

814c

411a

813a

1022a

Low level drain

813a

1021

8 km

River alluvium Fen peat Marine alluvium Raised moss peat Fen peat Marine alluvium (Boreal)

Series Name 411a 813a 814c 851a 1021 1022a

Association Name Evesham 1 Midelney NewChurch 2 Downholland 1 Turbary Moor Altcar 1

Lias clay and limestone Fig. 18.7 Soil associations on the Levels and Moors of Somerset and Avon. (Findlay et al. 1984.)

datum. Finally, the Evesham 1 association are calcareous soils developed on the Jurassic clays of the surrounding uplands. Alluvium contains complicated patterns of deposition that inevitably can be seen in the resulting soils (Figure 18.3). Typically, fluvial soils are silt dominated, the variation of texture reflects the pattern of deposition and origin of the sediment. Many typical chalk rivers, such as the Salisbury Avon, have floodplains dominated by Frome association soils (calcareous alluvial gley soils, equivalent to Food and Agriculture Organisation classification as epigleyic fluvisols). Where the major sources of alluvium are clay deposits, textures are finer. The Fladbury 3 association soils for example, on the Kent Stour floodplain, comprise pelo-alluvium gley soils (Jarvis et al. 1984), a finer alluvial soil, with topsoil texture clays or clay loams. It is used today for flood-meadows, grazing marsh (at the seaward end of the catchment) and (inappropriately) for intensive arable farming (Cook 2007a). From the seventeenth century, soils in alluvial valleys of the chalklands experienced the

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imposition of ‘floating’, that is irrigated water meadows (Cutting and Cummings 1999; Cutting et al. 2003) and the typical soil series is Frome. Storage in the chalk aquifer makes river flow relatively even and therefore more reliable for irrigation purposes than more ‘flashy’ catchments (Cook 2007b). Figure 18.8 shows a schematic for the soils of irrigated water meadows. Water travels along ‘carriages’ cut in the tops of ridges, and trickles down the sides to warm, oxygenate and fertilise the pasture topsoil. While Frome series soils dominate, there are also pockets of Gade (humose soils). These result from infill by vegetation in the drains between the channels and ridges (‘bedworks’). Viewed at a larger scale, however, pre-floating soil patterns inevitably occur. For example at Harnham Water Meadows, Salisbury, UK, humose Gade series dominate on lower lying floodplain areas, and appear to have been abandoned as functional water meadows before other (mineral soil) areas, probably because of the unsuitability of peaty soils to irrigation through very high infiltration rates (Cutting et al. 2003) and susceptibility to ‘poaching’ (trampling damage) by animal hooves.

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409

Wetland and Floodplain Soils

Hatch

Carrier

Fs

F

R

F F

F

Drain

G F

F G

Gravels Fig.18.8 Water meadows use ridges and channels to control flow and increase irrigation efficiency. The Frome series (‘F’, a calcareous alluvial gley) is dominant and a Frome shallow phase (Fs) is distinguished where gravel is encountered within 20 cm, a function of meadow construction. Racton series soils (R) are similar with non-calcareous subsoils. Gade series soils (G) are humose or peaty and calcareous within 40 cm and typically occur in topographic lows.

M A NA G E M E N T O F W E T L AN D S O ILS I N CH AN G E D F U T UR E Hydric soil processes and climate change Elevated atmospheric CO2 will have a significant impact on wetland soil processes owing to its indirect global effect on soil temperature. Temperature will affect the microbial processes that produce CO2, CH4 and N2O. The production of these gases are of particular relevance as they are all greenhouse gases and therefore could have a positive feedback on global warming. Soil production of CO2 is likely to increase because it is produced by heterotrophic aerobic respiration, however, there is discussion regarding the sensitivity of soil respiration to temperature (Davidson and Janssens 2006). Methane production is dependent on the balance between methanogenic versus methanotrophic respiration, both of which are highly dependent on the hydrological regime. The response of these microbial processes to

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temperature will depend on interacting environmental factors, that is, hydrology, nutrient supply and soil organic carbon. As mentioned above, the loss of N2O and N2 will be dependent on many interacting factors including nitrate, acidity, temperature and hydrology. Elevated CO2 potentially has a direct impact on plant productivity that may in turn affect soil processes via an increase in soil carbon, particularly within the rhizosphere as dissolved organic carbon. Primary productivity may even offset losses of CO2 from aerobic soil decomposition, although this will depend on hydrology and nutrient availability.

WET LAN D SOILS, AGR ICULT UR AL AN D ECOSY ST EM SER V ICES Human activities such as the burning of fossil fuels and land use change have caused recent changes in the world’s climate systems, and

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continued emissions of greenhouse gases are projected to result in further climate change (IPCC 2001a, 2007). These are expected to have impacts on wetland functioning and the supply of ecosystem goods and services. These goods and services include food, fibre (e.g. reeds, timber), clean water, stores or sinks for carbon and other nutrients, flood and storm control, groundwater recharge and discharge, pollution control, organic matter or sediment export, routes for animal and plant migration, and landscape and waterscape connectivity. With potentially dramatic changes in land use and climate likely within this century, the persistence of wetland ecosystems, and in some places their restoration, will be a necessary focus for conservation and ecosystem management. The value of wetlands to sequester or retain soil carbon must be balanced with the need to retain the numerous other functions and values that wetlands provide. For this reason, the value of wetlands to human society and the ways in which these ecosystems are managed are changing. No longer should peatlands, for example, be regarded as a source of fuel for burning, or as areas of potentially nutrient-rich land suitable for drainage and conversion for agriculture. Wetlands are critically important in global biogeochemical cycling, and climate change will have impacts on the biogeochemistry by affecting the hydrology, net primary production, respiration and decomposition rates, and C and nitrogen cycling in wetlands (Ramsar COP8 2002). Wetlands are correctly viewed as both a social and an environmental good. Nevertheless, while features such as high levels of productivity, biodiversity and flood alleviation are uncontroversial, there is a tendency to view wetlands in a negative fashion, for example as sinks for unwanted substances and providing attenuation of excess nutrients or barriers against pathogens (Cook and Kemp 1997). The real dichotomy lies in wetland use. The potential for wetland areas to be managed for biodiversity will meet many desirable objectives, including carbon sequestration. Historically, wetlands have been drained and used to achieve greater agricultural production. Today, they are increasingly studied as retainers

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of atmospheric carbon. With the world facing global climate change, ironically due largely to the burning of fossil fuels from ancient wetland deposits, their appropriate utilisation may be more relevant today than at any time in history. This is leading to the realisation that we must alter how we think about and manage wetlands and their resources. Even when they are not a major carbon sink, the restoration of wetland areas may promote the retention of significant amounts of carbon that, under different management, would have entered the atmosphere. In order to re-commence their natural function of carbon assimilation and storage, drained peats require re-hydration. This can involve alterations, including reversals, to the regional drainage patterns. While the Great Fen Project (http://www. greatfen.org.uk/) in East Anglia, UK, is aimed at improving biodiversity and restoring ‘landscape’, it is heartening that carbon sequestration is expected to follow with the revitalisation of the peat-forming process. This has attracted considerable technical attention in recent years Throughout the twentieth century, there was a threat to wetlands not just from agricultural drainage, but also from straightening and deepening main rivers, flood alleviation schemes, and over-abstraction of groundwater. The latter is particularly dangerous as it reduces the base flow of rivers. Supported in the UK by Land Drainage Acts and Water Acts, this culture of water management was to be changed by a new ‘conservation ethic’ in the 1980s (Cook 1998; Sheail 1999). Similar patterns of management can be found around the world where habitat loss, detrimental hydrological impacts and soil problems have emerged from inappropriate reclamation practices. There is a move therefore to slow or reverse this process. We may ask what wetland environments today can provide where production agriculture is to be curtailed or toned down. This enters the field of ecological restoration, beyond the scope of this chapter (see Section VI), but in outline, reversing the sequence in Table 18.1 is viewed as desirable and occurs in many places today. Choosing examples from Britain alone, we can identify the following.

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Wetland and Floodplain Soils Areas of arable reversion Tertiary wetlands in reclaimed coastal marshland may be returned to grazing marsh areas (Manchester et al. 1999). Sadly, the loss of many dykes and the reprofiling of ditches means the loss of many of the morphological features of reclamation, however, relict flora and fauna may become dominant once agro-chemicals are lost. Wicken Fen (http:// www.wicken.org.uk/), representing perhaps only 0.1% of British fenland, is the focal point for the planned phased restoration of hundreds of hectares of fenland in eastern England. Romney Marsh grazing marsh has been conserved, and there are plans to restore reed beds (http://www.kentbap. org.uk/) in the North Kent Marshes. Water Level Management Plans (WLMPs) are being produced to protect and enhance grazing marshes. Were there to be extensive reversion of floodplain soils from arable agriculture to grass, extensive opportunities for carbon sequestration would remain (Cook 2007a).

Water level management planning In the UK there has been a long history of draining the land, which has brought loss of habitat, amenity, flood detention storage, and ecosystem services. In certain river valleys, particularly the chalk river valleys of southern England, conservation imperatives (Sites of Special Scientific Interest (SSSIs), Special Areas of Conservation (SACs), and Ramsar sites) require the restoration of water levels, frequently depressed through drainage improvements both regionally and at the field scale, and through over-abstraction of groundwater. Water Level Management Plans were developed by the Environment Agency of England and Wales (Defra 2008) for operation until 2010. The objective is to restore water levels, especially in areas of particular conservation value. In Britain, these are SSSIs in particular. The Avon valley SAC comprises a range of measures to be brought to bear, including the introduction of barrages in rivers and channels, the closure of sluices, the blocking and abandonment of land drains and the relaxation of flood relief measures.

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Areas of coastal ‘managed retreat’ This is a management strategy not to be confused with land abandonment. Often sea defences are allowed to be, or are deliberately, breached in order to achieve a system of coastal defence that is economically sustainable in that the loss of agricultural land is planned. In this way, some secondary and tertiary wetland areas (Table 18.1) will move towards primary wetlands as salt marshes or estuarine mudflats. Areas in the UK earmarked for managed retreat include the North Kent Marshes and Sparkhayes Marsh, Somerset. Water meadows Water meadows, or more accurately ‘floated water meadows’ are commonplace but often derelict features of river valleys on the chalk areas of southern England. One reason for their decline was the great labour requirement in maintaining them throughout the agricultural year, another was finding alternative winter fodder for animals and modern methods of grass production, including artificial fertilisers and silage making. Decline has been significant for at least a century. In soil terms, water meadows were constructed using alluvial soils typically of the Frome association (Cutting et al. 2003). Re-working to produce bedwork ridges in the seventeenth century has left a legacy of topographic alteration of floodplain surfaces, but also of creating new depositions of soil and subsoil gravels (Figure 18.8). The uppermost 300 mm of soil need not be fully gleyed (i.e. grey or blue in colour, displaying reduced iron species) because the system is designed to allow infiltration of dissolved O2 together with the irrigation water entering the topsoil (Cutting et al. 2003). Below this, subsoil and gravel will generally be gleyed. Floating gardens Floating gardens are known from the Dal Lakes of India, and Mexico. The Dal Lake of Srinagar supports perhaps 100 settlements, and the floating gardens feed around 50% of the entire Kashmir

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Valley. Sadly, the depth of the lake is diminishing and its area shrinking due to the impact of horticulture and sewage effluent (http://www.boloji. com/wfs251htm). In Mexico, floating islands called chinampas are to be found dating back perhaps 1000 years, and still today demonstrate the potential for sustainable agriculture. The chinampas are formed by alternating layers of aquatic weeds, manure and earth packed inside rectangular cane frames rooted to the lake floor, and often surrounded by trees that can be used to secure the gardens to the lake bottom (http://www.planeta. com/ecotravel/mexico/df/xochililco.html). Understanding wetlands in the environment Functionality in wetlands needs to be understood. For example, the US Army Corps of Engineers promotes the quantification of functional aspects of wetlands based upon morphological considerations (Ashby 2002). What is proposed is a hydromorphic approach to wetland soils (see Section IV). Understanding wetland soils is a central aspect of environmental policy and conservation management. Examples include the ability of wetland soils to remove nitrate from subsurface flows, probably by a complicated balance of anaerobic denitrification, uptake by vegetation and assimilation in the soil biomass (White et al. 1998). Phosphorus too can be reduced in wetland environments and there is evidence that floating water meadows cause uptake of available (environmentally mobile) P and hence reduce the amount of P in suspended particulate matter or else dissolved as ‘orthophosphate’ in water returned to the river after irrigation. Predictions of the response of wetland soil processes to climate change is a long-term process, owing partly to the inaccuracy of global circulation models and partly to the short-term nature of experiments. Additionally, integrated management is required that encompasses the multitude of biogeochemical functions and values that wetlands provide. Many wetland areas in Europe have been destroyed and reduced over the last 250 years because they were not at the time regarded as of ‘value’ to mankind. Nevertheless,

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wetlands have always been important for biodiversity because of the vicinity of water and land, where many rare species can live and develop. In addition, wetlands have the function of being the ‘kidneys’ of a land, cleaning and regulating ground and surface waters. For example, management may be aimed at maintaining the hydrological regime of a wetland to support biodiversity, and may as a result increase methane emissions and thus offset carbon sequestration. Management to enhance soil denitrification of agricultural nitrate runoff may result in elevated nitrous oxide emissions that could offset CO2 fixed by plant productivity. Thus in many cases, a balance of management would be required that may not provide optimal control of specific soil or ecosystem processes. To protect biodiversity today in Europe and to guarantee clean water cycles and sources, largely means the protection of wetlands, but this cannot be done without reference to human activities. Therefore, there is a need to conciliate natural protection of wetlands with sustainable living opportunities for local populations. This is what ‘Integrated Management of Wetlands’ (http://www.wetlandsmanagement. org/) aims to achieve. Wetlands are drained for a range of reasons, but generally for agriculture. This is, and has been, the case with many peat areas. Rich ecosystems have been replaced in many instances by extremely acid sulphate soils and infertile organic soils. The original systems are destroyed forever. Knowledge about soil development and readiness to its use could have avoided these major catastrophes. We are asking not only: ‘Why waste the worlds wet places?’ (Maltby 1986), but even worse: ‘Why completely destroy rich, highly productive wetland soils to obtain unproductive land?’ R EFER EN CES Allen J.R.L. 1970. Physical Processes of Sedimentation. An Introduction. George Allen and Unwin, London, 248 pp. Aselmann I. and Crutzen P.J. 1989. Global distribution of natural freshwater wetlands and rice paddies,

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19 The Role of Buffer Zones for Agricultural Runoff M A R T IN S .A. B L ACKWELL 1 , D AVI D . V. H O G AN 2 , GIL L E S P INAY 3 AND ED WARD MALTBY 4 1Biogeochemistry

of Soils and Water Group, North Wyke Research, Okehampton, UK 2Environmental Consultant, Exeter, Devon, UK 3School of Geography, Earth and Environmental Sciences, University of Birmingham, Birmingham, UK 4Institute for Sustainable Water, Integrated Management, and Ecosystem Research, University of Liverpool, Liverpool, UK

IN T R O D U CT ION The twentieth century witnessed major changes in agricultural practices, not least in the intensification of the use of nitrogen (N) and phosphorus (P) fertilisers to promote crop productivity. Likewise, increasingly intensive livestock production resulted in the need to dispose of large quantities of animal waste and, consequently, the spreading on farmland of manure and slurry rich in N and P compounds has intensified (Addiscott et al. 1991). These practices regularly result in the export of large quantities of nutrients from the application areas in runoff water or as leachates. A contemporaneous aspect of agricultural development has been intensive land drainage, which has included the conversion of riparian wetlands to agricultural use. This has resulted not only in an increase in the potential source of nutrients subject to transfer to the wider environment, but also the degradation and loss of ecosystems that are capable of reducing or buffering the flux of nutrients from terrestrial to aquatic environments (Tiner 1984; Hollis and Jones 1991;

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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Lake et al. 2000; Brinson and Malvárez 2002; Zedler and Kercher 2005). The water quality buffering potential of wetlands was first reported in detail in the early 1970s. Subsequently, much material has been published describing the numerous benefits that can be gained from wetlands, often supported by data demonstrating their high efficiency in improving water quality (Peterjohn and Correll 1984; Dugan 1990; Davies and Claridge 1993; Haycock et al. 1997; Kronvang et al. 2003). This has resulted in a ‘Swiss Army Knife’ attitude to wetland buffer zones, where it is assumed that they are multi-purpose, capable of buffering any amount and type of pollution. While buffer zones are often highly effective at improving and maintaining water quality in surface water bodies, there are limitations to their capabilities. In order to optimise this role alongside other environmental benefits, it is vital to understand the mechanisms that enable a wetland buffer zone to function as a water quality regulator. This chapter discusses the ability of freshwater wetlands to function as buffer zones, regulating water quality by storing and transforming nutrients through the performance of a variety of physical, chemical and biological processes. ‘Buffer zone’ is a generic term referring to naturally or semi-naturally vegetated areas situated

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between agricultural land and a surface water body. These areas can be effective in protecting the water body from harmful impacts such as high nutrient, pesticide or sediment loadings resulting from land use practices (Blackwell et al. 1999). The degree to which this protection is provided depends on a number of factors including the size, location, hydrology, vegetation and soil type of the buffer zone (Leeds-Harrison et al. 1996; Dosskey et al. 1997; Polyakov et al. 2005), as well as the nature of the threats to the water body. It is important to understand that the terms buffer zone and wetland are not synonymous. The role of a buffer zone is one of protection of a given environmental structure, and consequently some buffers comprise land that is dry, well drained and lacking the characteristics of a wetland. In addition, neither buffer zones nor wetlands are confined to riparian areas but may be located at some distance from the river channel or other water body, including land on adjoining slopes, but usually in association with some type of hydrological pathway for pollutants (e.g. a ditch or area of overland flow). Riparian areas themselves often comprise a complex association of wet and dry types of land. Permeable land on floodplains, which includes levees formed adjacent to the river channel, cannot be regarded strictly as wetlands by most definitions (e.g. presence of hydromorphic soils or characteristic hydrophytic vegetation) but can perform certain functions comparable with wetlands if they are susceptible to flooding. Moreover, conditions required to provide effective buffering for one nutrient may be ineffective for another. A wetland’s buffering ability is strongly dependent on intrinsic biogeochemical processes that vary among different nutrients and their compounds. Additionally, the hydrogeomorphic context (hydrology and landform) of the wetland must be considered. For example, if a wetland is not linked hydrologically to an agriculturally-derived nutrient source, it will be unable to function as a buffer zone for the removal of agricultural nutrients, though modification of hydrological pathways may provide that opportunity. However, careful consideration must be given to hydrological

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modifications of this type, and their consequent impact on the wetland, in order to avoid the risks of both hydrological and chemical overloading, and the consequent degradation of vegetation and soils. In this chapter we highlight the buffering functions of river marginal wetlands and focus on the factors controlling the efficiency and ability of wetlands to act as buffer zones against diffuse fluxes of N and P. We also discuss approaches to the management and implementation of buffer zones at both local and catchment scales. River marginal wetlands can be described as floodplain ecosystems where a high water table or inundation by surface water is frequent. This may include adjacent valley slopes where conditions are similar. Soils should be sufficiently waterlogged for their development as hydromorphic soils, supporting hydrophytic vegetation. Availability of N and P has been identified by many researchers as critical in controlling productivity in wetlands and aquatic systems (Reader 1978; Nixon and Lee 1986; Knight 1992; Rejmánková 2005); in excess these nutrients can even cause ecosystem degradation. The biogeochemical dynamics of N and P respectively in wetlands are described in detail by White and Reddy (Chapter 9) and Richardson and Vaithiyanathan (Chapter 10), while the cycling of these nutrients and their interaction is discussed by Verhoeven (Chapter 12 ).

MECHAN ISMS OF N UT R IEN T CON T R OL BY WET LAN D BUFFER Z ON ES A wetland buffer zone can assist with water quality improvement by limiting nutrient fluxes to aquatic ecosystems in a variety of ways. If located adjacent to a water body, the physical presence of, for example, trees, shrubs and other rough vegetation can restrict livestock access thereby reducing impacts such as accelerated bank erosion and direct nutrient inputs. These physical obstructions may also prevent the direct application of fertilisers, pesticides and other pollutants to the buffer areas themselves by precluding access for machinery, thus extending the distance between

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The Role of Buffer Zones for Agricultural Runoff applications and a receiving water body and providing vegetation capable of intercepting airborne particles. However, control by wetland buffer zones of nutrients in runoff originating from agricultural land relies largely on the dynamics of N and P biogeochemical transformation processes. The main processes affecting N dynamics are denitrification, nitrification, plant uptake, fixation, adsorption, bacterial assimilation and volatilisation (see White and Reddy, Chapter 9). The key processes affecting P dynamics are plant uptake, bacterial assimilation, adsorption, sedimentation, precipitation and re-mobilisation under anaerobic conditions (see Richardson and Vaithiyanathan, Chapter 10). The generalised spatial relationships of the main processes of N and P transformation in a typical wetland soil are shown in Figure 19.1. The dominant processes are determined largely by the nature of the wetland, which in turn depends mainly upon the intrinsic variables of

Harvesting of vegetation (N and P)

hydrology, soil, vegetation and size. Extrinsic variables such as climate and local agricultural practices are also significant, the former influencing factors such as hydrology and rate of biochemical processes, and the latter affecting the quantity and type of nutrient fluxes in the environment. Figure 19.2 illustrates the generalised relationships among these variables. Only intrinsic variables are discussed in this chapter. Hydrology Three main aspects of the hydrology of a wetland buffer zone control its ability to transform and store nutrients: • hydrological linkage with the wider catchment; • internal hydrological regime; • internal hydrological pathways. While the hydrological linkages between wetlands and sources of nutrients in the wider

Gaseous emission (N)

Water

Fixation (N)

Volatilisation (N)

Storage (N and P)

Sedimentation (P)

Adsorption (N and P)

Aerobic soil layer

Nitrification (N)

Anaerobic soil layer

Oxidation by plant roots Plant uptake (N and P)

Mineralisation (N and P) Denitrification (N)

Bacterial assimilation (N and P)

Anaerobic release of P

Fig. 19.1 Generalised diagram showing spatial relationships of key nutrient transformation processes in a typical wetland. (Adapted from Mitsch and Gosselink 1993.)

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Climate

Soil type

Topography

Geology/soils

Catchment land use

Hydrological connectivity

Nutrient loading

Landscape location

Vegetation

Hydrology

Extrinsic (distal) factors

Size

Intrinsic (proximal) factors

Buffer zone management

Wetland buffer zone functioning Fig. 19.2 Generalised schematic diagram showing factors influencing wetland buffer zone functioning.

catchment largely control the opportunity a wetland has to process nutrients (discussed below), the effectiveness of that processing actually depends on the wetland’s internal hydrological regime and hydrological pathways (Ross 1995; Maltby et al. 1996; Burt et al. 1999; Pinay et al. 1999; Hill et al. 2000). In turn, these are influenced by many factors, including soil type, location in the catchment and geomorphology. For optimal removal of most nutrients, the water table is required to be at or near the soil surface for long periods. These sites are characterised by hydric soils and hydrophytic vegetation (Jacinthe et al. 2000). The soil environment is predominantly anaerobic, although aerobic zones often occur, especially near the surface. The spatial and temporal relationships and interactions between oxidised and reduced conditions can be critically important for carrying out both biotic and abiotic nutrient transformation. Examples of the relationships among hydrology,

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wetland type and wetland location in the catchment of the River Tamar, south-west England, are illustrated in Figure 19.3. These factors have been used to group wetlands on a functional basis as shown in this figure, with the primary division being between slope (S) and floodplain (F) wetland units. Details of the wetland functional units presented in Figure 19.3 (e.g. S1–S4 and F1–F3) and their potential to perform processes influencing water quality are given in Table 19.1, and discussed in more detail below. Spatial differences in aerobicity (oxygen status) within a wetland soil are usually associated with the presence of plant roots, soil pore spaces or surface water–soil interfaces, and are characteristically revealed in various patterns of mottling in mineral soil material. Nitrification (the oxidation of ammonium to nitrate (NO3) by nitrifying bacteria in the soil), often takes place within aerobic zones. This can result in the establishment of a

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The Role of Buffer Zones for Agricultural Runoff

1. Wetlands on footslopes only (headwaters)

(a)

(b)

WBZ S2 or S3

S4

S1

(c)

WBZ S1

River channel

2. Wetlands on footslope and floodplain (middle catchment)

(a) S2 or S3

S4

F2

(a) dry F3

F2

F1

River channel KEY S1–S4 : Slope units F1–F3 : Floodplain units WBZ: Wetland Buffer Zones : Main hydrological flow paths : Intermittent or minor hydrological flow paths

River channel

WBZ dry

F2

F2

River channel

(b)

WBZ

S1

(c)

WBZ

River channel

3. Wetlands on floodplain only (lower catchment)

dry

River channel

(b)

WBZ

WBZ

(c)

WBZ dry

F2

River channel

F1

River channel

no WBZ dry

F1

River channel

Fig. 19.3 Examples of generalised hydrology and geomorphic setting of river marginal wetland buffer zones and relationships to wetland functional units.

NO3 concentration gradient between the aerobic and anaerobic sites, promoting the diffusion of NO3 and consequently a regular supply of NO3 for denitrifying bacteria in the anaerobic parts of the soil. Denitrification involves the reduction of oxidised forms of nitrogen (N), in particular NO3, to gaseous forms, primarily di-nitrogen (N2) and nitrous oxide (N2O), by facultative anaerobic bacteria (Knowles 1981; Bowden 1987; Tiedje 1988). Anoxic conditions are essential for the process, since they stimulate denitrifying organisms to use NO3 as an electron acceptor in the absence of free oxygen, thus causing NO3 to progressively be reduced to elemental N (Clément et al. 2003; Hefting et al. 2003). The particular importance of this process in terms of nutrient removal is

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that it involves total export of N from the hydrological system to the atmosphere as gas, rather than temporary, though sometimes long-term, storage by other processes such as plant uptake or organic matter accumulation in the soil. The ability of wetlands to remove NO3 via denitrification has been reported by many researchers including Lowrance et al. (1984), Peterjohn and Correll (1984), Ambus and Lowrance (1991), Cooper (1990), Simmons et al. (1992), Pinay and Labroue (1986), Jordan et al. (1993) and Blackwell et al. (1999), all of whom report reductions of 75% or more in NO3 concentrations in surface or groundwater compared with the original concentrations in water discharging into the wetlands studied.

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Geomorphic position

Floodplain – level and/or elevation

Floodplain – level/ depression

Floodplain – backland/ depression

Base of footslope (0–2°), including tributary valley floors lacking alluvium

Unit name

F1

F2

F3

S1

Ditches, water tracks

Ditches, oxbows, abandoned channels

Ditches, slacks

Occasional depressions, slacks

Other features

High groundwater and some surface flow

Permanently high groundwater, seepage and surface flow inputs; flooding

Flooding; high groundwater table

Occasional flooding; seasonal groundwater in subsoil

Hydrodynamics

Slowly permeable stagnogleys

Humic and alluvial gley soils

Alluvial gley soils

Permeable, brown alluvial soils

Soils

Tall herb fen, willow carr, patches of flood grass along water tracks, some improved to rush pasture

Tall herb fen, willow carr, flood pasture

Rush pasture, willow scrub and woodland

Mesophile dry grassland, scrub and woodland

Vegetation

Plant uptake of nutrients Denitrification Adsorption of N and P

Plant uptake of nutrients Sediment retention Denitrification Storage of organic matter Adsorption of N and P

Plant uptake of nutrients Sediment retention Denitrification Storage of organic matter Adsorption of N and P

Plant uptake of nutrients Sediment retention

Key processes potentially occurring in unit that influence water quality

Table19.1 Classification of the wetland functional units occurring in the River Tamar catchment, Devon, UK, based on geomorphology, hydrodynamics, soil type and vegetation, and the key processes occurring within them that influence water quality.

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Footslope (0–2°)

Footslope (0–2°)

Footslope (2–5°)

Backslope (>5°)

S2

S3

S4

S5

Irregular surface, slumping

Ditches

Peat mounds, floating vegetation mats

Peat mounds

Seepage and overland flow

High groundwater and rain-fed with soils of low permeability, occasional seepage

Strong groundwater discharge fringing alluvium, on tributary valley bottoms, lower slopes and in valley heads

Strong groundwater discharge fringing alluvium, on lower slopes and in valley heads

Gleys of varied permeability

Slowly permeable stagnogleys and stagnohumic gleys

Peat

Peat

Rush pasture and willow scrub

Humid grassland, fen meadow, wet heath (includes Culm grassland), willow, improved land with rush pasture

Tussock sedge, willow, flood grass

Tussocky Molina and bog

Low nutrient input means potential to perform nutrient retention/ removal is low

Plant uptake of nutrients Denitrification Storage of organic matter Adsorption of N and P

Low nutrient input means potential to perform nutrient retention/ removal is low

Low nutrient input means potential to perform nutrient retention/ removal is low

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total NO3 loss occurring in the riparian zone. This disproportionate role of the organic soils in depleting NO3 was attributed to the fact that their low-lying position meant that 37–81% of groundwater flowed through them on its way to the stream, and that they were anaerobic with a high carbon content, and consequently had high denitrification potential. Blackwell (1997) and Baker and Maltby (1995) reported denitrification rates from what have been termed ‘zones of enhanced denitrification’. These were discrete areas of overland flow on a degraded ditch system draining agricultural land and passing through a wetland in the catchment of the River Torridge, Devon, UK (see Figure 19.4). Denitrification rates as high as 29.3 mg N m−2 d−1 were reported by Blackwell (1997), while rates as low as 0.28 mg N m−2 d−1 were measured in the wetland immediately adjacent to the overland flow zones. The high denitrification rates contributed to reductions regularly in excess of 90% of NO3

Sep-95

Jul-95

Aug-95

Jun-95

May-95

Jul-94

T $

Aug-94

River Torridge

Surface water bodies

NO3 −N, kg

Improved pasture N

9 8 7 6 5 4 3 2 1 0

Sep-94

Nitrate-N loads

Overland flow zones 100 m

Jun-94

The coupling of nitrification and denitrification can result in the export of large quantities of N from systems where there is no apparent external NO3 source, or where the reduced N compound, ammonium, is abundant (Reddy et al. 1978, 1989). It may result from water table fluctuations (Patrick and Wyatt 1964), whereby NO3 accumulates under aerobic conditions when the water table is lowered, and denitrification occurs when the water table subsequently rises and anaerobic conditions are re-established. In these circumstances, the amount of N lost from a system will depend largely upon the frequency of water table fluctuation, which can range from daily to seasonal, depending on the hydrology of the wetland. The influence of NO3 availability and hydrological pathways through a wetland was demonstrated by Cooper (1990), who reported that riparian organic soils, occupying only 12% of a headwater stream’s riparian zone in New Zealand, were responsible for 56–100% of the

Month

T $

T $

T $

Jul-95

Aug-95

Sep-95

Jul-95

Aug-95

Sep-95

Jun-95

Sep-94

May-95

Aug-94

Jul-94

Jun-94

$ T

Removal Efficiency

Nitrate removal efficiency 100 90 80 70 60 50 40 30 20 10 0

Month

Jun-95

Sep-94

May-95

Aug-94

Jul-94

Jun-94

$ T

NO3- −N, kg

Nitrate-N loads

9 8 7 6 5 4 3 2 1 0

Month

Fig. 19.4 Kismeldon Meadows: an example of alternative wetland buffer zone locations for the removal of nitrate from agricultural runoff. (From Maltby et al. 2000.)

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The Role of Buffer Zones for Agricultural Runoff loads in the ditch system. These results indicate the importance of opportunity for interaction in soil that has a high denitrifying potential and a NO3 supply, in order that NO3 removal may take place. The NO3 supply in this case was in the form of runoff from grazed grassland. More recently, Syversen (2004) has investigated methods of ensuring that the construction of ditches that intercept subsurface drains, and disperse drainage water across wetlands do not cause the by-passing of wetland buffer zones. Rates reported for P removal in natural wetland buffer zones generally are not as great as those for N removal. The important processes of P retention are precipitation, adsorption, sedimentation and plant uptake, however, the internal hydrology of a wetland can actually reduce its capacity for P removal. Phosphorus bound to ferric iron under aerobic conditions can be desorbed and re-mobilised if a change to anaerobic conditions and a consequent reduction in redox potential occurs, as the ferric iron is used as an electron acceptor in the place of oxygen (Caraco et al. 1989; Reddy and D’Angelo 1994; Lamers et al. 1998). This may result from a rise in the water table or when precipitates become buried by sedimentation, and can lead to a net increase in P loss from a wetland buffer zone. Particulate P is reported in some cases to be stored in large quantities through sedimentation, adsorption, and plant uptake (Braskerud 2002; Venterink et al. 2003). Sedimentation can occur when the flow velocity of runoff is reduced as it enters a wetland with ponded water (Kadlec and Knight 1996). Mander et al. (1991) reported retention of P entering a natural wetland in Estonia of between 27% and 88%, while Jenssen et al. (1993) measured retention rates of total P of 98% in an artificial wetland in southern Norway. However, if flows are concentrated within a wetland, deposition rates decline and erosion may occur, resulting in the buffer zone acting as a source rather than a sink of particulate nutrients. This can arise as a result of changing morphology of a buffer zone due to the deposition of large amounts of sediment (Jordan et al. 1993). Generally, it is reported that none of the P transformation

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processes within wetlands result in large scale, complete export of P from the hydrological system. Most is merely stored over a range of time periods in sediment or vegetation. In constructed wetlands, different results are reported, largely because conditions can be controlled and vegetation is frequently harvested as part of their management, resulting in nutrient removal (see above and Kadlec, Chapter 20). However, in some circumstances P potentially can be exported from a wetland system via the production and emission of phosphine gas (PH3) (Dèvai and Delaune 1995). Under the reduced conditions typically prevalent in wetlands, Dèvai et al. (1988) reported that oxidised P compounds were reduced to phosphine, which was subsequently emitted to the atmosphere at a rate of 1.7 g P m−2 per year (a−1). To date there has been very little research into this process, which requires further investigation in order to determine if it is significant in terms of quantity of P exported, and the fate of gaseous P. Glindemann et al. (2005) have provided a useful review of research into the evolution of phosphine in a range of sediments. Soil type Hydrological characteristics of a soil are determined by hydraulic conductivity, moisture retention and pathways of water movement (Boorman et al. 1995). In general terms, wetlands are maintained either by a persistent or fluctuating high groundwater table in permeable or low lying soils, or by the presence of material of low permeability within the soil or substrate, which prevents the downward percolation of rainfall and causes the formation of a perched water table. Many soil properties, such as bulk density and organic matter, are susceptible to change as a result of land use (Hall et al. 1977) or the succession of vegetation, while the property least likely to be affected in this way is texture. Sandy soils generally are permeable and tend to drain readily, though some can be susceptible to compaction and the development of plough pans when in arable use, allowing standing water to develop following wet spells. In low-lying locations lacking

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artificial drainage, a naturally high groundwater table leads to the development of intense waterlogging. The organic matter status of the soil is important to processes such as denitrification, which requires a carbon supply. The presence of humified organic material can promote adsorption of P and other nutrient ions. Organic (peat) soils develop in permanently waterlogged sites and often possess limited N-cycling activity, especially under acidic conditions. In naturally well-drained soils, water will tend to pass down through the soil profile readily, except where occurring on steep slopes, during temporary periods of waterlogging, or where poor timing or inappropriate land-work has led to the formation of layers of low permeability (pans). In these cases, water will either flow away as runoff or remain ponded on the surface. Sediment that is transported in runoff that ultimately infiltrates soil is liable to become trapped in the soil pores; the size of the particles deposited reflecting the size of the pores (i.e. the smaller the pore size the smaller the particles that can be trapped). This process has been found to be particularly effective at trapping finer particles to which the majority of nutrients usually are adsorbed (Dillaha et al. 1988; Metcalf and Eddy 1991; Hosokawa and Horie 1992). The sustainability of this process in buffer zones is not clear, but is undoubtedly finite, becoming less efficient with time. Infiltration can be promoted by the presence of roots that help the development of soil structure. The infiltration of runoff may also provide greater opportunities for other nutrient removal processes to occur, such as adsorption and plant uptake. Vegetation The productivity of biomass in wetlands can be high (Johnston 1991) and so, therefore, will be the potential capacity for nutrient removal by plant uptake. Since plant growth is strongly seasonal, so are the processes of nutrient removal and storage by plant uptake. During dormant periods, the senescence and decomposition of above-ground plant material results in the release of nutrients. Klopatek (1978) reported rates of N and P

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uptake in riverine marshes in Wisconsin, USA, of 20.8 g N m−2 a−1 and 4.3 g P m−2 a−1 respectively, though only 26% of N and 38% of P were retained within the vegetation, with the rest being returned to the system in leaching and litterfall. The amount of nutrient uptake by plants depends upon their growth rates, age and stage of development. Storage time also varies among plant types, generally being longest in woody vegetation, although the nutrient concentrations found in plants of this type generally are low, averaging 0.4% N and 0.01% P dry weight (Johnston 1991). In a mature wetland forest this can equate to 95 g N m−2 and 4.3 g P m−2 (Schlesinger 1978). If nutrients are abundant then ‘luxury’ uptake may occur, resulting in increases in the concentration of nutrients in the vegetation (e.g. Feller et al. 2002; Newman et al. 2004). Kadlec and Knight (1996) report an increase of N in the leaves of cattails from an average of 1.6–2.0% in zones of N-enriched discharge. In overall terms, the quantity of nutrients entering into long-term storage in woody vegetation is small. It is estimated that the amount of P allocated annually to woody growth equates to only 0.1 g m−2 a−1 (Johnston et al. 1996). Nutrients taken up by non-woody plants may enter into long-term storage if the vegetation becomes incorporated into peat. Typically peaty topsoils contain 2.5% N and 0.2% P based on dry matter (Brady and Weil 1996). However, as peat accumulates very slowly, annual retention rates are likely to be relatively small. The harvesting of biomass is one method of nutrient removal from a wetland system, but again only very small amounts are likely to be removed. Herskowitz (1986) reported that only 2.5% of the total P removed from a wetland was the result of harvesting biomass, besides which, harvesting is often difficult and labour intensive, making it an inefficient procedure for P-removal. In addition to nutrient removal by uptake, vegetation may act as a sediment filter, depending on its structure and density (Dillaha et al. 1989; Braskerud 2001). Larger soil particles, to which nutrients can be adsorbed, can be trapped within vegetation litter layers or among stems and leaves. Vegetation may also act to slow the flow

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The Role of Buffer Zones for Agricultural Runoff of runoff, and the consequent reduction in energy results in the further deposition of suspended particles. Densely packed, finely structured vegetation such as grass habitats offer the most resistance to flows at shallow depth (i.e. they have the greatest roughness coefficients), but their impact can be less when flows are at greater depth. The slower the flow, the finer the particles deposited. The sorting of sediment deposits will in turn influence the N-cycling in floodplain soils because generally the finer the sediment the higher the denitrification activity (Pinay et al. 1995). Vegetation also indirectly contributes to the regulation of N fluxes by fuelling the denitrification process through the supply of organic carbon (Haycock et al. 1993). Vegetation type has also been reported as influencing nutrient removal. For example, Haycock and Pinay (1993) reported that a riparian zone of poplar was more effective than a meadow riparian zone for N removal from shallow groundwater, while others have found no significant differences (Gilliam et al. 1986; Groffman et al. 1991; Clément 2001). In a european research programme evaluating N control in various agricultural landscapes across different climates (NICOLAS; Pinay and Burt 2001), results indicated no significant difference in N-removal capability between forested and riparian meadow sites. Size and shape Various techniques have been devised for determining minimal or optimal buffer zone widths and areas, most of which are applicable to wetlands. For example, simple ratio methods exist based on comparisons of the area of wetland to the area of stream catchment; ratios of 1 : 10–1 : 50 wetland area to catchment area have been suggested for significant reduction in pollutant transfer to occur (Williams and Nicks 1993; UusiKämppä and Yläranta 1996). This method does not use site-specific data, can be adapted for different regions, and is simple to apply once the ratio has been determined. Another simple approach is that devised by Trimble and Sartz

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(1957), which relates wetland buffer zone area to catchment slope and catchment area, optimal buffer width being positively related to both these variables. However, neither of these approaches considers buffer zone shape, which can be important for increasing buffer zone effectiveness (e.g. Gergel et al. 2005). Complex models for determining buffer zone width have been devised, such as the Riparian Ecosystem Management Model (REMM) (Lowrance et al. 1998), the Chemicals, Runoff and Erosion from Agricultural Management Systems Model (CREAMS) (Kinsel 1980), and the Water Erosion Prediction Project (WEPP) model (Flanagan and Nearing 1995), all of which require considerable amounts of data relating to fundamental hydrologic and erosion mechanics, making their application impractical in many cases. A summary of recommended widths of buffer zones for the performance of different functions derived from a review of buffer zone literature is shown in Figure 19.5. A wetland buffer zone usually can be considered as occupying the land between two system boundaries: terrestrial-wetland and wetlandaquatic systems. In wetlands where the main source of water is diffuse shallow groundwater, it is often the wetland-terrestrial interface that is the functionally most active zone, particularly with regard to NO3 removal. Nitrate concentrations are sometimes depleted so rapidly within this zone, often over a distance of a few metres, that the rest of the wetland does not have the opportunity to remove it despite having the potential for N removal. As a consequence, it is the length of interface between wetland and upslope sources of nutrient-rich runoff as opposed to overall size (i.e. the shape of the buffer zone), that can be most important in determining the capacity of a wetland to buffer surface water bodies from nutrient-rich runoff. Haycock and Pinay (1993), reporting from Oxfordshire, UK, indicated almost complete removal of NO3 from groundwater within the first 5 m on entering a riparian buffer zone, with concentrations typically falling from 9.0 mg NO3-N L−1 to less than 0.5 mg NO3-N L−1 over

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Air quality Recreation Instream Bankside conservation

Function

Bank stability Max

Flood defence

Mean Min

Water quality – pesticides/herbicide Water quality – sediment retention Water quality – stream Water quality – BOD Water quality – nutrients 0

50

100

150

200

250

300

Width, m Fig. 19.5 Summary of published maximum, mean and minimum effective widths of buffer zones in relation to different functions. Based on Haycock and Muscutt 1995.

this distance. Peterjohn and Correll (1984) found greatest NO3 reduction within the first 17 m of a 50 m buffer zone, while Cooper (1990) reported a reduction in NO3 concentrations in soil water from 0.64 mg NO3-N L−1 to 0.01 mg NO3-N L−1 within the first few metres of a 15 m wide buffer zone. Cooper et al. (1986) found riparian strips 16 m wide were effective for NO3 removal, with mean NO3 concentrations falling from 7.6 to 0.2 mg NO3-N L−1 across the buffer zone. Lowrance (1992) reported decreases in NO3 concentrations by factors of between 7 and 10 in the first 10 m of a riparian woodland, while Pinay and Dècamps (1988) reported similar findings in a 50 m wide buffer zone, and Simmons et al. (1992) found removal of NO3 in

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groundwater in excess of 80% in what they term the ‘transition zone’ (effectively the upslope interface) between a wetland and well-drained soils upslope. The optimal location for NO3 removal at the upslope edge of a riparian buffer zone receiving water from the hill slopes is explained in most cases by the fact that, at this interface, the combination of high NO3 concentrations in the runoff, high carbon concentrations in the soil (essential as a respiratory substrate for the denitrifying bacteria), and anaerobic conditions resulting from a typically elevated water table at the wetland interface, provide optimal conditions for denitrification to occur. Haycock and Pinay (1993) found that, as discharge and NO3 concentration

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The Role of Buffer Zones for Agricultural Runoff in groundwater upslope of the wetland increased, higher NO3 concentrations did not penetrate much further into the wetland, as suggested by Cooper (1990) and Pinay and Dècamps (1988). Instead, the zone of NO3 removal, or active interface, was found to extend further upslope, as the water table rose in that direction. The importance of the active interface has implications for the location of wetlands in terms of both their optimum efficiency for nutrient removal from agricultural runoff, and their economics. For example, it would be more effective for nutrient removal to convert an area of agricultural land into wetland in the form of long thin strips, close to surface water bodies, than to create one single wetland block of equivalent area. If 10 m wide strips were distributed around the farm, ten times more upslope wetland-agricultural land interface would be created than with a single block. If, however, wildlife habitat creation is the main objective, then a more compact area may be preferable. In addition to riparian wetlands along low order streams, there are many other landscape features that can provide the opportunity for establishment of a dry–wet interface at which NO3 removal may occur. These include features such as ditches, slope wetlands and hedges with ditches (Caubel-Forget and Grimaldi 2000). The capacity of a wetland to remove nutrients from runoff will depend largely on its length of interface with the upland input of N in relation to the quantity of water passing through it. The relationships among wetland shape, wetland size, discharge, catchment area and potential of wetland for improving water quality can be complex. Other factors, including catchment soil type, vegetation, land use, topography and climate will also affect the quality and quantity of runoff that a wetland is able to process. Yet practical factors sometimes govern the size of a buffer zone, for example, economic considerations related to the amount of land required to accommodate the landscape features that can perform nutrient buffering activities. Consequently, buffer zones often are constrained by the fact that they should not incur high costs or loss of income to a landowner

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or manager, while still performing the desired functions to a sufficiently high degree. Wetland management The management requirements for wetland buffer zones depend largely upon the type of functions they are expected to perform. The nutrient removal function is enhanced by the periodic removal of biomass and any sediment deposited, though this is not always essential. Processes such as denitrification and, to a lesser extent, phosphine production (see above) can convert nutrients to gases which are emitted to the atmosphere, making wetland buffer zones direct nutrient sinks, rather than just stores. One potentially economically viable option for harvesting vegetation from a wetland buffer zone is coppicing. It may be possible to plant newly established buffer zones with rapidly growing trees that have a high capacity for nutrient uptake. Consequently, income may be gained from the production of timber, and simultaneously high rates of nutrient removal may be achieved, with actual export from the system as opposed to long-term storage. Typically, species of willow and alder are used for this purpose in temperate climates, and coppicing is practiced to maximise timber production. Coppiced woodland is one of the richest wildlife habitats, supporting a wide range of woodland plants and animals. Rotational cutting strategies maintain diverse habitat, but also ensure a constant and high level of nutrient uptake. They can also provide good economic returns, Nix (2009) estimating an income of £177 ha−1 a−1. Another common practice is the cutting of grassland buffer zones for hay or silage, which again removes nutrients stored in vegetation, while at the same time providing a direct economic benefit to the landowner. The economics of this practice are optimised by using nutrients in runoff from fields upslope as the source of fertiliser, as opposed to direct application of farm waste or mineral fertilizer. Another economically viable approach to harvesting wetland buffer zone vegetation is by grazing them during dry

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periods in summer, when grass growth in drier pastures has declined, and livestock require supplementary food supplies. However, this option is less desirable as it can result in re-mobilisation of nutrients in livestock excreta, and the direct loss of biodiversity (Foote and Rice Hornung 2005). Loss of species diversity may also result from the increased productivity of a wetland as it strips nutrients from eutrophicated water. This is particularly important in wetlands where productivity is N-limited, such as herbaceous fens (Koerselman and Verhoeven 1992). Careful management, however, can result in improved habitat type, and allow wetland buffer zones to be stocked with game birds such as woodcock and pheasant, providing an additional source of income. In riparian buffer zones, angling may also provide economic benefits, but management is required to maintain access and provide the riparian and channel habitats required. Sediment will be trapped and prevented from further downstream movement if it is deposited in wet depressions. However, if a buffer zone’s capacity for storage is reached, sediment will be no longer removed and the buffer zone will become ineffective. Sediment can be excavated or dredged and redistributed over farmland as a source of nutrients, but in many wetlands this is impracticable or uneconomic (Dillaha and Inamdar 1997). To date there has been little research into this aspect of buffer zones, as this technique is relatively new, and only now are the problems of buffer zone capacity and life expectancy becoming evident. In some wetlands, it is possible to manage the water table to promote nutrient transformation processes, for example by creating fluctuating water tables that promote coupled nitrification– denitrification and also provide wildlife habitat. In the N cycle, some gaseous intermediary compounds and by-products of the denitrification and nitrification processes, such as NO and N2O, are important since they are greenhouse gases (Wayne 1993). There is controversy over the respective contributions of nitrification and denitrification to atmospheric N2O (Groffman et al. 2000), but

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Jacinthe et al. (2000) suggest that the best management option to sustain high NO3 removal rates and to reduce the proportion of N2O in the emitted gases, is to maintain a high water table for a prolonged period in the most biologicallyactive portion of the soil profile.

WET LAN DS IN T HE DR AIN AGE BASIN CON T EX T Location in the catchment The location of a wetland in the landscape can be the crucial factor in determining its effectiveness as a buffer zone (Johnston et al. 1990; Johnston 1991). In the past, the assessment of wetlands at the catchment scale has focused on the total area of wetlands present (Omernik et al. 1981; Osborne and Kovacic 1993; Hayakawaa et al. 2006) and, for functions such as habitat provision and sediment storage capacity, this approach is useful. However, the geomorphic context (i.e. relative position in catchment and spatial relationship to other landscape features) was found to be most important in determining differences in the N-buffering capacities of riparian sites (Pinay and Burt 2001), while spatial relationships can vary seasonally with drainage network expansion, resulting in the by-passing of ‘active’ areas of riparian buffers (see below, and Wigington et al. 2005). However, the examples above demonstrate that for the purposes of assessing the nutrient removal capacity of wetlands within a whole catchment, this approach alone is inadequate because the amount of wetland–terrestrial ecosystem interface, which is a crucially important factor for this function, is not taken into account. The shape and orientation of wetlands in relation to slopes and surface water bodies is important as it affects the amount of upslope, ‘active’ interface present within a catchment. In terms of nutrient (especially NO3) removal, active interfaces are those at which there is a flux of nutrients from the terrestrial to the wetland ecosystem, and generally include only interfaces on the upslope boundary of wetlands. One exception

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The Role of Buffer Zones for Agricultural Runoff to this is on large floodplains during periods of sudden high flow in otherwise dry periods, when the water table in the floodplain is low. In such circumstances, there may be flow, and consequently a flux of nutrients, from a river to a wetland (Burt 1997; Wigington et al. 2005). Moreover, there are both anthropogenic and natural reasons for hydrological by-passing of a buffer zone, which render them ineffective for water quality improvement. The installation of drains and ditches that intercept runoff and discharge it directly to surface water bodies is one of the main anthropogenic causes of buffer zone ineffectiveness. Natural hydrological pathways in some floodplains can also result in buffer zone inefficiency, particularly those bordering rivers of high order (i.e. the lower reaches of a catchment) where flows can be at depth via substrate gravels rather than through the surface and upper soil layers of the wetland (Burt 1997). In these cases, both artificially drained floodplains and those with highly permeable substrates result in the rapid passage of hill slope runoff directly to the river, linking the channel efficiently with the slope. These are some of the reasons why Johnston et al. (1990) reported that wetlands bordering streams of low order (i.e. in the upper reaches or headwaters of a catchment) are more effective as buffers than those bordering streams of high order. Consequently, the establishment or conservation of wetlands along streams of low order is most likely to result in the greatest protection of water quality. Towards functional assessment of wetland buffering capacity at the catchment level Models for the design of buffer zones, such as REMM (Lowrance et al. 1998) and CREAMS (Kinsel 1980), are useful at the local scale, but if the overall health of a river is to be improved, a broader, more strategic, catchment-wide approach to buffer zone protection, restoration and creation is necessary (Viaud et al. 2004; Lake 2005; Merot et al. 2006). For many reasons, a blanket approach to buffer zone implementation, whereby all riverbanks are protected by buffer zones, is not

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practical and not necessarily beneficial. Economic constraints will nearly always apply, and a balance between farming practices, economics and river protection has to be found. It is important that natural processes in rivers be allowed to continue for channel and floodplain development, and the overall health of the river. There is a danger that riverbanks will be over-protected, preventing natural river erosion and change. Therefore buffer zones must be established that provide what Gardiner and Perala-Gardiner (1997) refer to as ‘space of liberty’ for streams. Lateral movement of the channel is permitted, and this, despite creating local instability, results in a more stable, sustainable river system overall. Issues such as this must be addressed at a catchment scale by determining where and by how much intervention is required. In Europe, the Water Framework Directive (2000/60/EC) relies on a participatory and multi-disciplinary approach to address water quality and quantity issues at the catchment scale (Clarke et al. 2003; Newson and Chalk 2004; Thenail and Baudry 2005; Bateman et al. 2006). Simple models have been developed for predicting the effects of land use change on catchment nutrient losses, such as the export coefficient modelling approach devised by Johnes (1996), and refinements of it (e.g. Beven et al. 2005). This approach calculates the total nutrient load delivered to a water body as a sum of individual loads exported from each nutrient source in its catchment. It allows the scaling-up of plot scale studies to provide a practical catchment scale approach to land use change impact assessment. In this context, consideration of the impact of wetland buffer zones on nutrient fluxes in terms of their spatial relationship to other types of land use could prove useful for assessing the impact of wetland buffer zone creation schemes. A functional classification of wetlands based on their geomorphic position and hydrology can form a useful first stage in a wider assessment of buffer zone functional capacity and catchment planning. As part of the Tamar 2000 SUPPORT Project (Sustainable Practices Project On the River Tamar; Maltby et al. 1997), a wetland

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inventory was established in which a series of mappable units was identified, comprising readily identifiable and distinguishable wetlands, each assessed in terms of geomorphic position, hydrology, and consequent ability to perform a range of functions delivering environmental benefits. This was determined using Wetland Functional Assessment Procedures (Maltby 2008; see Maltby et al., Chapter 23) developed under the FAEWE (Functional Analysis of European Wetland Ecosystems) (Maltby et al. 1996) and PROTOWET (Procedural Operationalisation of Techniques for the Functional Analysis of European Wetland Ecosystems) projects, funded by the environment programme of the European Commission, DGXII Science, Research and Development. They comprise a set of procedures intended to enable nonspecialists to predict the functional capability of

wetland systems as an aid to resolving dilemmas in planning, land management and impact assessment. The wetland functional units identified in the Tamar 2000 SUPPORT Project are separated broadly into those located on floodplains (coded F) and those on slopes (coded S), and then subsequently into functional groups based on their generalised hydrology, soil type and nutrient status. A description of each unit is given in Table 19.1, and a diagrammatic representation of the typical location of each functional unit is shown in Figures 19.6a and b. Each functional unit has the potential to perform a different range of functions, or similar functions to different degrees. Such a functional classification can assist with buffer zone planning at a catchment scale by providing an indication of the types of functions they would be performing, many of

(a)

F1 F2

F3 F1 F2

F3

Peat

(b) #

S1 S2

F2

S4

S5 F1

Fig. 19.6 Diagrammatic representations of the typical location of different wetland types found in the Tamar catchment, Devon, UK.

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The Role of Buffer Zones for Agricultural Runoff which would be useful in the context of a buffer zone (e.g. denitrification, sediment retention and plant uptake). Consequently, this assists with assessment of current buffering capacity for different pollutants and identification of optimal location for the establishment of new buffer zones to deal with specific pollution issues. For example, if a catchment had a particular problem with NO3 pollution, wetland units with the ability to perform the function of denitrification (e.g. F2, F3, S1 and S4) could be targeted preferentially for restoration. The determination of where buffer zones can be established for effective functioning requires consideration of many factors. Baudry (1997) investigated how farm structure and farming systems can affect the practicalities of buffer zone establishment and presence at a landscape scale in northern France. It was found that the larger the relative area of riparian zone within a farm unit, the more likely it was to be used intensively, making it less likely that any riparian vegetation would remain intact for use as buffer zones, or that such areas would be created. Also, the type of farming system influences the requirement for land, which consequently affects riparian land use. These factors indicate that policies regarding wetland buffer zones must be designed within a context of landscape and farming systems, and give consideration to areas outside the riparian zone.

CON CL US ION Research has shown that wetlands can provide a low maintenance opportunity for reducing the input of nutrients and other agriculturally derived pollutants to surface and shallow groundwaters. The effectiveness of wetlands as buffer zones largely depends on their location in the landscape with regard to connectivity with hydrological pathways and their spatial relationship with water bodies, both of which affect the opportunity to buffer nutrient loads. There is consensus among wetland researchers that the impact of buffer zones is greatest along low order streams,

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for the control of both dissolved and particulate forms of nutrients. Processes such as denitrification, that result in the complete export of nutrients from a system, are the most beneficial in terms of nutrient control in the long term. Other processes such as plant uptake and sediment deposition can be effective in the short term, but require management techniques such as harvesting and dredging to ensure permanent export. The majority of processes acting in wetland buffer zones result in only the storage of nutrients, which always have the potential for re-mobilisation. As a buffer zone matures, or if its hydrology changes, it can become a nutrient source rather than a sink. Temporal changes in wetland buffer zone functioning require further investigation. Optimal buffer zone width is a contentious subject, and one that is difficult to resolve. The reason for this is that optimal size will depend on many variables, not least the nature of the wetland, the functions it is to perform, and economic constraints. For many nutrient retention and export functions, it is not the overall area or width of buffer zone that is important, but its configuration within the landscape, particularly the length of interface between the wetland and agricultural land upslope, which determines the optimal rate of many nutrient removal processes. Numerous models exist for the design of buffer zones, but generally they require large amounts of site-specific data that are not usually readily available. In most cases, it is appropriate to take a general approach to wetland buffer zone implementation developed from the results of a large number of individual case studies. These include simple ratio methods relating buffer zone width to buffer zone catchment size and slope, and the determination of standard widths for riparian or field margin buffer zones. These can be adjusted to accommodate local needs including economic, environmental and spatial variables. Functional approaches to wetland buffer zones are being developed and provide mechanisms for scaling-up from field scale studies to the planning of whole catchment strategies.

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Because of the temporal and spatial uncertainty in the functioning of wetlands for nutrient removal, wetland buffer zones should not be relied upon as the sole method for controlling nutrient inputs to water bodies. They can be a very effective factor in preventing nutrients entering surface waters, but they should be used in conjunction with other in-field and off-field Best Practice management procedures for fertiliser, land and water management as part of a holistic, catchment-wide approach to environmental management.

ACK N OW L E D G E M E N T S The authors acknowledge the contribution to the development of the overall understanding of buffer zones made by the EU funded NICOLAS Project, ENV4-CT97-0395, FAEWE Project (ECDG XII STEP-CT90-0084) and PROTOWET Project (EC DG XII ENV4-CT95-0060). Thank you to Victoria Cook for assistance in the production of the diagrams presented in this chapter.

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quality. Critical Reviews in Environmental Control 21, 491–565. Johnston C.A., Detenbeck N.E. and Niemi G.J. 1990. The cumulative effect of wetlands on stream water quality. Critical Reviews in Environmental Control 21, 491–565. Johnston C.A., Schubauer-Berigan J.P. and Bridgham S.D. 1996. The potential role of riverine wetlands as buffer zones. In: Haycock N.E., Burt T.P., Goulding K.W.T. and Pinay G. (editors), Buffer Zones: Their Processes and Potential in Water Protection. Quest Environmental, Harpenden, UK, pp. 155–170. Jordan T.E., Correll D.L. and Weller D.E. 1993. Nutrient interception by a riparian forest receiving inputs from adjacent cropland. Journal of Environmental Quality 22, 467–473. Kadlec R.H. and Knight R.L. 1996. Treatment Wetlands. Lewis Publishers, CRC Press, Boca Raton, FL, 893 pp. Kinsel W.G. 1980. CREAMS: A Field Scale Model for Chemicals, Runoff, and Erosion From Agricultural Management Systems. US Department of Agriculture, Conservation Report No. 26, 640 pp. Klopatek J.M. 1978. Nutrient dynamics of freshwater riverine marshes and the role of emergent macrophytes. In: Good R.E., Whigham D.F. and Simpson R.L. (editors), Freshwater Wetlands Ecological Processes and Management Potential. Academic Press, New York, pp. 195–216. Knight R.L. 1992. Ancillary benefits and potential problems with the use of wetlands for nonpoint source pollution control. Ecological Engineering 1, 97–113. Knowles R. 1981. Denitrification. In: Paul E.A. and Ladd J.M. (editors), Soil Biochemistry, Vol. 5. Marcel Decker, New York, pp. 323–369. Koerselman W. and Verhoeven J.T.A. 1992. Nutrient dynamics in mires of various status nutrient inputs and outputs and the internal nutrient cycle. In: Verhoeven J.T.A. (editor), Fens and Bogs in the Netherlands: Vegetation, History, Nutrient Dynamics and Conservation. Kluwer Academic Publishers, Dordrecht, pp. 397–432. Kronvang B., Laubel A., Larsen S.E., Andersen H.E. and Djurhuus J. 2003. Buffer zones as a sink for sediment and phosphorus between the field and stream: Danish field experiences. Paper from Diffuse Pollution Conference Dublin 2003, available at: http://www. ucd.ie/dipcon/docs/theme03/theme03_02.PDF, last accessed on 26 January 2009. Lake P.S. 2005. Perturbation, restoration and seeking ecological sustainability in Australian flowing waters. Hydrobiologia 552(1), 109–120.

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The Role of Buffer Zones for Agricultural Runoff Lake P.S., Palmer M.A., Biro P., Cole J., Covich A.P., Dahm C., Gibert J., Goedkoop W., Martens K. and Verhoeven J. 2000. Global change and the biodiversity of freshwater ecosystems: impacts on linkages between above-sediment and sediment biota. BioScience 50, 1099–1107. Lamers L.P.M., Van Roozendaal S.M.E. and Roelofs J.G.M. 1998. Acidification of freshwater wetlands combined effects of non-airborne sulfur pollution and desiccation. Water, Air and Soil Pollution 105, 95–106. Leeds-Harrison P.B., Quinton J.N., Walker M.J., Harrison K.S., Tyrrel S.F., Morris J.M. and Harrod T. 1996. Buffer Zones in Headwater Catchments. Report on MAFF/English Nature Buffer Zone Project CSA 2285. Cranfield University, Silsoe, UK. Lowrance R.R. 1992. Groundwater nitrate and denitrification in a coastal plain riparian forest. Journal of Environmental Quality 21, 401–405. Lowrance R., Altier L.S., Williams R.G., Inamdar S.P., Bosch D.D., Sheridan J.M., Thomas D.L. and Hubbard R.K. 1998. The Riparian Ecosystem Management Model Simulator for Ecological Processes in Riparian Zones. Proceedings of the First Federal Interagency Hydrologic Modelling Conference, Las Vegas, NV, April 1998, pp. 1.81–1.88. Lowrance R.R., Todd R.L. and Asmussen L.E. 1984. Nutrient cycling in an agricultural watershed I. Phreatic movement. Journal of Environmental Quality 13, 22–27. Maltby E. (editor) 2008. Functional Assessment of Wetlands: Towards Evaluation of Ecosystem Services. Woodhead Publishing, Cambridge, UK. Maltby E., Blackwell M.S.A. and Hogan D.V. 1997. Wetlands in the Landscape – linking policy to function. In: Soderqvist T. (editor), Wetlands: Landscape and Institutional Perspectives. Proceedings of the 4th Workshop of the Global Wetlands Economic Network (GWEN). Beijer Occasional Paper Series, Beijer International Institute of Economics, The Royal Swedish Academy of Sciences, Stockholm, Sweden, 16–17 November 1997, pp. 265–291. Maltby E., Hogan D.V. and McInnes R.J. 1996. Functional Analysis of European Wetland Ecosystems – Phase I (FAEWE). Ecosystems Research Report 18. Office for Official Publications of the European Communities, Luxembourg, 448 pp. Maltby E., Blackwell M.S.A. and Baker C. 2000. Linking Wetland Science to Policy: Meeting the challenge with special reference to water quality issues. In: Balazs E.,

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Galante E., Lynch J.M., Schepers J.S., Toutant J-P., Werner D. and Werry P.A.T.J. (editors), Biological Resource Management: Connecting Science and Policy. Springer-Verlag, Berlin, Heidelberg, New York, pp. 291–308. Mander Ü., Matt O. and Nugin U. 1991. Perspectives on vegetated shoal, ponds and ditches as extensive outdoor systems of wastewater treatment in Estonia. In: Etnier C. and Guterstam B. (editors), Ecological Engineering for Wastewater Treatment. Proceedings of the International Conference at Stensund Folk College, Sweden, Mars 24–28, pp. 271–282. Merot P., Hubert-Moy L., Gascuel-Odoux C., Clément B., Durand P., Baudry J. and Thenail C. 2006. A method for improving the management of controversial wetland. Environmental Management 37(2), 258–270. Metcalf and Eddy Inc. 1991. Wastewater Engineering, Treatment, Disposal and Reuse. (3rd edition). Revised by G. Tchobanoglous and F.L. Burton. McGraw-Hill, New York. Mitsch W.J. and Gosselink J.G. 2000. Wetlands (3rd edition). Van Nostrand Reinhold, New York. Newman S., McCormick P.V., Shi Li Miao, Laing J.A., Kennedy W.C. and O’Dell M.B. 2004. The effect of phosphorus enrichment on the nutrient status of a northern Everglades slough. Wetlands Ecology and Management 12(2), 63–79. Newson M. and Chalk E. 2004. Environmental capital: an information core to public participation in strategic and operational decisions – the example of river ‘Best Practice’ projects. Journal of Environmental Planning and Management 47(6), 899–920. Nix J. 2009. Farm Management Pocketbook (39th edition). Imperial College London, Wye Campus. The Andersons Centre, Melton Mowbray, UK, pp. 268. Nixon S.W. and Lee V. 1986. Wetlands and Water Quality. A Regional Review of Recent Research in the United States on the Role of Freshwater and Saltwater Wetlands as Sources, Sinks and Transformers of Nitrogen, Phosphorus and Various Heavy Metals. Final Report Prepared for the Department of the Army, US Army Corps of Engineers. Waterways Experiment Stations, Technical Report Y-86-2. Omernik J.M., Abernathy A.R. and Male L.M. 1981. Stream nutrient levels and proximity of agricultural watershed observations on the role of a riparian forest. Journal of Soil and Water Conservation 36(4), 227–231.

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Osborne L.L. and Kovacic D.A. 1993. Riparian vegetated buffer strips in water quality restoration and stream management. Freshwater Biology 29, 243–258. Patrick W.H. and Wyatt R. 1964. Soil nitrogen loss as a result of alternate submergence and drying. Soil Science Society of America Proceedings 28, 647–653. Peterjohn W.T. and Correll D.L. 1984. Nutrient Dynamics in an agricultural watershed observations on the role of a riparian forest. Ecology 65, 1466–1475. Pinay G. and Burt T.P. 2001. Nitrogen Control by Landscape Structures in Agricultural Environments (NICOLAS), Executive Summary. ECDGXII, Environment and Climate ENV4-CT97-0395. Web page http://www.aopv55.dsl.pipex.com/nicolas/nicolas.htm, last accessed on 26 January 2008. Pinay G. and Dècamps H. 1988. The role of riparian woods in regulating nitrogen fluxes between the alluvial aquifer and surface water A conceptual model. Regulated Rivers: Research and Management 2, 507–516. Pinay G., Décamps H. and Naiman R.J. 1999. The spiralling concept and nitrogen cycling in large river floodplain soils. Archiv fur Hydrobiologie Suppl. 11(3), 281–291. Pinay G. and Labroue L. 1986. Une station d’épuration naturelle des nitrates transportés par les nappes alluvialles: l’aulnaie glutineuse. Comptes Rendus de l’Academie des Sciences de Paris, III, 302 III: 629–632. Pinay G., Ruffinoni C. and Fabre A. 1995. Nitrogen cycling in two riparian forest soils under different geomorphic conditions. Biogeochemistry 30(1), 9–29. Polyakov V., Fares A. and Ryder M.H. 2005. Precision riparian buffers for the control of nonpoint source pollutant loading into surface water: a review. Environmental Reviews 13(3), 129–144. Reader R.J. 1978. Primary Production in Northern Bog Marshes. In: Good R.E., Whigham D.F. and Simpson R.L. (editors), Freshwater Wetlands: Ecological Processes and Management Potential. Academic Press, New York, pp. 53–62. Reddy K.R. and D’Angelo E.M. 1994. Soil processes regulating water quality in wetlands. In: Mitsch W. (editor), Global Wetlands Old World and New. Elsevier, Amsterdam, pp. 309–324. Reddy K.R., Patrick W.H. Jr. and Lindau C.W. 1989. Nitrification-denitrification at the plant root sediment interface in wetlands. Limnology and Oceanography 34, 1004–1013.

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Reddy K.R., Patrick W.H. Jr. and Phillips R.E. 1978. Ammonium diffusion as a factor in nitrogen loss from flooded soils. Soil Science Society of America Journal 40(4), 528–533. Rejmánková E. 2005. Nutrient resorption in wetland macrophytes: comparison across several regions of different nutrient status. New Phytologist 167(2), 471–482. Ross S.M. 1995. Overview of the hydrochemistry and solute processes in British wetlands. In: Hughes J.M.R. and Heathwaite A.L. (editors), Hydrology and Hydrochemistry of British Wetlands. Wiley, London, pp. 133–181. Schlesinger W.H. 1978. Community structure, dynamics and nutrient cycling in the Okefenokee cypress swamp-forest. Ecological Monographs 48, 43–65. Simmons R.C., Gold A.J. and Groffman P.M. 1992. Nitrate dynamics in riparian forests groundwater studies. Journal of Environmental Quality 21, 659–665. Syversen N. 2004. Filtering of Nutrients and Pesticides through Vegetated Buffer Root Zones (In Norwegian). Jordforsk-report 118/04. Norwegian Centre for Soil and Environmental Research, Aas, Norway, 26. Thenail C. and Baudry J. 2005. Farm riparian land use and management: driving factors and tensions between technical and ecological functions. Environmental Management 36(5), 640–653. Tiedje J.M. 1988. Ecology of denitrification and dissimilatory reduction of nitrate to ammonia. In: Zehnder A.J.B. (editor), Biology of Anaerobic Microorganisms. John Wiley, New York, pp. 179–244. Tiner R.W. 1984. Wetlands of the United States: Current Status and Trends. US Department of the Interior, Fish and Wildlife Service, US Government Printing Office, Washington, 59 pp. Trimble G.R. and Sartz R.S. 1957. How far from a stream should a logging road be located? Journal of Forestry 55, 339–341. Uusi-Kämppä J. and Yläranta T. 1996. Effect of buffer strips on controlling soil erosion and nutrient losses in southern Finland. In: Mulamoottil G., Warner B.G. and McBean E.A. (editors), Wetlands Environmental Gradients, Boundaries and Buffers. CRC, Lewis Publishers, New York, pp. 221–235. Venterink H.O., Wiegman F., Van der Lee G.E.M. and Vermaat J.E. 2003. Role of Active Floodplains for Nutrient Retention in the River Rhine. Journal of Environmental Quality 32, 1430–1435.

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20 Wetlands for Contaminant and Wastewater Treatment R OB E RT H . KAD LEC 1,2 1University 2Wetland

of Michigan, Ann Arbor, USA Management Services, Chelsea, USA

IN T R O D U CT ION Types and applications Treatment wetlands are constructed ecosystems designed primarily to enhance physical, chemical and biochemical processes, with the goal of reducing specific contaminants to acceptable levels. Water quality improvement is the primary driving force for selection and sizing. Other wetland functions and benefits, such as flood control, habitat and human uses, are relegated to secondary importance. Nevertheless, such secondary benefits often accompany wetland water treatment. Natural wetlands have long been known to provide water quality improvement when they become the recipients of contaminated flows. In past and present times, pre-existing wetlands have been the receiving waters for imperfectly treated wastewaters. Studies have shown that such discharges are improved upon passage through the wetlands (Fetter et al. 1978; Mudroch and Capobianco 1979). These older a posteriori studies deal with wetlands as receiving waters, with no conscious attempt to employ the water quality improvement functions of the ecosystems. The added water and nutrients have been observed to cause impacts on the soils and biota of natural wetlands, to the extreme of complete alteration The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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of the species abundance and composition (Kadlec and Bevis 1990). In the decades 1950–1970, laboratory studies of the pollutant reduction potential of wetland plants were undertaken in Germany, targeting phenol and other industrial chemicals (Seidel 1966). In the 1970s, engineered treatment wetland projects were first undertaken in the United States. Some involved engineering of natural wetlands (Tilton and Kadlec 1979; Ewel and Odum 1984); others were constructed on previous upland (James and Bogaert 1989; Litchfield 1993). One of the first reports of the use of subsurface flow (SSF), gravel bed wetlands in the USA was by Spangler et al. (1976). Many types of natural wetlands have served as prototypes for treatment wetlands (see Whigham, Chapter 2). These are grouped under the designation of free water surface (FWS) systems (Figure 20.1). Emphasis may be placed on floating, submerged or emergent vegetation, but some degree of hybridisation is inevitable. Constructed treatment wetlands are most frequently dominated by emergent macrophytes, such as Phragmites spp., Schoenoplectus spp. and Typha spp., because these species are tolerant of a wide variety of water chemistry and hydrology. The hydrologic discriminators of hydroperiod and water regime are often limited by the desire to maximise the treatment potential of a given land area. To that end, the hydroperiod of a treatment wetland is almost always set to be 100%. The desire to minimise water level adjustment operations leads to relatively constant water depths, typically set within the range of 10–50 cm. This

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Wetlands for Contaminant and Wastewater Treatment Treatment wetlands Paper SSF

FWS

Subsurface flow

Free water surface

Petroleum

HF

VF

FAP

SAV

EM

Horizontal flow

Vertical flow

Floating aquatic plants

Submerged aquatic vegetation

Emergent macrophytes

SMB

ATS

Submerged macrophyte beds

Algal turf scrubbers

Fig. 20.1 Classification of treatment wetlands. A variety of plant types and species may be utilised in any one category, such as a mix of herbaceous and woody emergent macrophytes.

range corresponds to the water tolerance of many emergent wetland macrophytes. A second general family of treatment wetlands is based on a water saturated bed of porous media. A variety of soils, sands and gravels have been employed, in either vertical or horizontal flow (Figure 20.1). These might be more appropriately termed ‘gravel bed hydroponic systems’, because there is no true analogue in the realm of natural wetlands. In common nomenclature, these are SSF wetlands. The plants are emergent macrophytes, of the same species used in FWS wetlands. The purposes of the media are basically two-fold: a microbial substrate for treatment and prevention of exposure of the contaminated water. The range of usage of wetland treatment technology is large, and still expanding (Figure 20.2). The largest number and the longest history belong to the class that treats municipal and domestic wastewaters. Europe has thousands of SSF wetlands treating settled sewage (Vymazal et al. 1998; Vymazal and Kropfelova 2008), and there are hundreds of FWS wetlands in North America further treating secondary and tertiary municipal effluents (Kadlec and Wallace 2008). A closely related

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Urban Drinking stormwater water

Domestic tertiary

Crop runoff Animal wastewater Remediation

Food

Domestic secondary

Metal mine drainage Acid mine drainage Landfill leachate Sludge consolidation

Domestic primary

Treatment wetlands Fig. 20.2 Applications of treatment wetlands. Any one category, such as landfill leachate, may span a wide range of water chemistry and treatment goals.

but newer employment is in the treatment of animal wastes (CH2M Hill and Payne Engineering 1997). Hundreds of FWS wetlands have been built to treat acid mine drainage in North America and Europe (Weider 1989; Younger et al. 2002). A variety of industrial applications have emerged in the past decade, including petroleum (Knight et al. 1999), petrochemical and paper production (NCASI 2004). Urban and agricultural stormwaters are particularly amenable to wetland treatment, and are fast becoming an important branch of the technology. Suggested methodologies are available for field runoff (USDA 1991) and urban settings (Schueler 1992; Lampe and Grizzard 1999; Carleton et al. 2001). Also fast growing is the utilisation of passive wetlands to treat contaminated groundwaters, which may arise from landfill and metal mine dump leaching, and old industrial spills and other sources of groundwater contamination (Mulamoottil et al. 1998; Sobelewski 1999; Means and Hinchee 2000; Nehring and Brauning 2002; Kadlec 2003; Nivala et al. 2007). The technology is in a phase of compound growth, with proliferation of use areas compounding the growth in each individual type of application.

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Treatment wetland technology has a very large literature spanning the last three decades. Entire books and monographs have been produced dealing exclusively with the subject (Kadlec and Knight 1996; Vymazal et al. 1998; IWA Macrophyte Group 2000; Kadlec and Wallace 2008; Vymazal and Kropfelova 2008). There are many volumes of Water Science and Technology devoted to papers from international conferences (19 : 10, 29 : 4, 32 : 3, 35 : 5, 40 : 3, 44 : 11/12, 48 : 5, 51 : 9, 56 : 3). Conferences have produced numerous other edited volumes (Reddy and Smith 1987; Hammer 1989; Cooper and Findlater. 1990; Moshiri 1993; Vymazal 2001, 2003, 2005b, 2008; Mander et al. 2003; Mander and Jenssen 2003a,b). This very extensive science and engineering literature base is among the best for any water pollution control technology. Mechanisms Complex wetland ecosystems offer a range of chemical processing mechanisms that transform, degrade or store nearly any incoming substances (Figure 20.3). Waterborne constituents may be transferred to the wetland sediments by settling,

sorption, chemical precipitation or accretion of residuals from the biogeochemical cycles. Incoming particulates find a quiescent environment in which to settle, and effective trapping in the wetland litter. Those particulates often carry other associated chemicals, and thus solids trapping removes those as well. Organic chemicals often partition preferentially to the organic sediments found in wetlands, acting as a natural carbon filter. Wetlands create reactive substances that are available to chemically bind incoming pollutants. An important example of this is the formation of sulphide in the anaerobic zones of the wetland, which can then act to precipitate a number of divalent metal cations, and to drive autotrophic denitrification. The undecomposed fraction of dead biological material, together with its fractions of several potential pollutants, also contributes to storage in wetland solids. Microorganisms, including bacteria and algae, mediate a number of processes. For instance, bacteria mediate both nitrification and denitrification, which combine to reduce total nitrogen in the wetland environment. Algae mediate the formation of calcium minerals, which eliminate

Biomass storage Gasification Volatilisation

Export

Biochemical transformation

Deposition sorption

Residuals accretion

Fig. 20.3 A variety of wetland processes act to reduce an incoming contaminant load. However, by-products of chemical conversion and the biogeochemical cycle may be released to the water, creating potential increases in some water quality indices.

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Wetlands for Contaminant and Wastewater Treatment contaminants by precipitation and co-precipitation. Wetland plants, plant detritus and sediments form the attachment media for those organisms – very few are found in water suspension. FWS wetlands are often operated within a large expanse of water and relatively long detention times. These are ideal conditions for the volatilisation of chemicals that tend to partition to air. Ammonia nitrogen, light hydrocarbons and gaseous metabolites – carbon dioxide, oxygen and nitrogen – have important wetland pathways into and out of the water and rhizosphere. The reader is referred to the several chapters of Section II for more details of wetland biogeochemical functioning. Operational strategies Constructed treatment wetlands differ from natural wetlands in the possibility of controlled water additions and water levels, as well as the potential for chemical additions and modifications of soils and water. There has consequently evolved a spectrum of operational procedures, designed to augment natural processes. At one end of that spectrum are totally passive systems, in which water flows and depths are regulated only by the character of the conveyance structures. The only processes that are considered in design are those which are sustainable over long time periods (decades). No credit is given for temporary storages in soils or biota. No upkeep activities are contemplated, other than maintenance of the physical integrity of berms and structures. These ‘Type A’ treatment wetlands enjoy the lowest operational costs of any version of the technology. They are the most numerous of surface flow treatment wetlands. It is human nature to desire more control over the performance of the system than are afforded by Type A wetlands. The possibilities for such controls are limited to manipulations of the soils, water, biota and environment of the wetland. Two general classifications of operational augmentation exist: those that are sustainable indefinitely (Type B), and those that require the de-commissioning and rebuilding of the wetland

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(Type C). Type B augmentation may be physical or chemical. For instance, the thermal environment of the wetland can be altered by putting it in a greenhouse, or by running heating pipes through the soils. The amount of oxygen reaching the rhizosphere may be increased by either bubbling air via perforated tubing, or by intermittent draining and re-flooding of a gravel bed wetland. The variations on these additions are limited only by human ingenuity, and are sometimes covered by patents. Harvest of wetland vegetation is also an option under the Type B designation. However, experience has shown that the recovery potential is typically small, and the degree of difficulty is very large (Vymazal 2005a; Kadlec and Wallace 2008). Only above ground plant parts can be accessed without total disruption, and these are not usually the most important storage location. The storage capacity of wetland soils and substrates can have a significant lifetime of temporary storage for some substances, such as metals and phosphorus. If the initial wetland substrate has a large binding capacity, it can be a significant number of years before that capacity is reached. These Type C wetlands are to be rebuilt at the end of the useful life of the substrate. During their lifetime, they may also provide sustainable functions, such as nitrogen removal. It may be economically feasible to replace media that have a long life. For instance, preliminary research indicates that expanded clays may be capable of providing many years of phosphorus removal (Zhu et al. 1997), but commonly available sands and calcitic materials are not (Brix et al. 2001). It is becoming more apparent that SSF wetlands will clog over some few years, thus requiring some form of drastic maintenance, such as excavation, cleaning and re-establishment (Cooper et al. 2008).

ST R UCT UR E AN D FUN CT ION Configuration Constructed treatment wetland systems are comprised of a number of shallow basins or

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cells, connected in series and parallel combinations. Small wetland cells are frequently built with rectangular geometry, with aspect ratios in the range 1 : 1–25 : 1 length to width. That geometry is the easiest to draw and build, but is not the most aesthetically pleasing. If the ancillary benefit of green space for human enjoyment is added to the design criteria, then it is possible to incorporate more elegant contours (Figure 20.4). Parallel cell trains allow for off-line maintenance activities while operating the balance of the system. Serial cell arrangement creates a favourable hydraulic mode, with minimal possibilities for short-circuiting.

Each horizontal flow cell is contained within a berm of approximately one metre in height, with side slopes typically of an aspect 1 : 3 height to width (Figure 20.5). The cell is sometimes lined with an impermeable material if it is necessary to prevent leakage of water to an underlying aquifer. Liners may be polymeric sheets, or layers of native or bentonite clay. However, if there is no threat to regional groundwater, some leakage may be allowed, which is then subjected to soil aquifer treatment. The cells are then filled with approximately 30 cm of topsoil as a rooting medium if the FWS option is being employed. The basins are filled with gravel, sand or soil media

Eskilstuna River

E20

Park

Treatment Plant

Folkestaleden

Fig. 20.4 Layout of the Ekeby treatment wetlands at Eskilstuna, Sweden. Water passes through five cells in parallel to intermediate collection, and then through three cells in parallel to the river. Gentle curves, deep zones and islands accentuate the aesthetic appearance of the system, and enhance its efficiency. See Plate 20.4 for colour version of this image.

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445

Outlet deep zone Live

FWS

Standing dead

Inlet deep zone

Detritus

Roots and rhizomes

Impervious liner

Hydric soils

SSF

Liner bedding

Live Rock-filled inlet deep zone

Underlying strata

Rock-filled outlet deep zone

Standing dead Detritus

Roots and rhizomes

Impervious liner

Water-filled gravel bed

Dry gravel

Liner bedding

Underlying strata

Fig. 20.5 Structural components of FWS and SSF constructed treatment wetlands. The FWS system is filled with a layer of hydric soil and overlain by shallow water. The SSF system is filled with gravel in the central portion and rock in the inlet and outlet zones. Both have outlet depth control. See Plate 20.5 for colour version of this image.

to prevent exposure of the water being treated if the SSF option is desired. Vertical flow cells contain a layer of permeable rock on top of the liner, in which are embedded perforated drain pipes to convey the collected water to downstream treatment or reuse. The basin is then filled with gravel of the correct draining characteristics.

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Water is brought into a horizontal flow cell via control structures and a distribution system. Inflow structures are typically level-lip or V-notch weirs. Distribution may be accomplished with either a piping manifold with adjustable outlets, or a cross ditch that can be filled with coarse rock, in the case of a SSF wetland, or left open in

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the case of a FWS wetland. FWS cells also may contain open water areas at internal locations. Collection is typically accomplished by a cross ditch that can be filled with coarse rock (SSF), or left open (FWS). Level control is provided at the outlet by means of a weir or swivelling pipe. Soils, media and plants The soils of a FWS wetland provide several zones of differing redox potential, and thus different chemistries. There is typically a layered vertical profile, with mildly reducing conditions just below the soil–water interface, progressing to methanogenic conditions in or below the root zone (c. 30 cm) (Reddy and D’Angelo 1994). In addition, gas interchange via plant roots leads to oxic zones very near the roots and anoxic zones at very small distances away. The amount of radial root oxygen loss is enough to protect the root, and depends on the oxygen demand in the nearby soils. The carbon content of treatment wetland soils develops toward high percentages, 30–40%, due to the accretion of detritus from both above and below ground plant parts, as well as the remains of other biota. This accretion is in the range of 0.1–2.0 cm a−1 for wetlands (Mitsch and Gosselink 2000). The porous media that fill the SSF wetlands vary from fine sands and soils, to relatively large stones in some systems. The choice of size is a compromise between the desire to maximise surface area within the media, favoured by small size, and the desire to have adequate hydraulic conductivity to keep the water below grade, favoured by large size. Cooper et al. (1996) recommend sizes in the range 3–12 mm. These media are most often mineral materials, either mined or river gravels, or crushed and graded stone. The porosities of the placed media are in the range of 35–40%, thus removing over half of the available wetland volume from water occupancy. The most obvious external feature of treatment wetlands is the vegetation. The green expanse, punctuated by patches of open water in FWS systems, conveys a warmer feeling than the concrete and steel of higher technologies. It is ironic that the green portion of the wetland often plays a less

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important role in treatment than do other components of the ecosystem: the litter, sediments, soils, bacteria, fungi, algae, roots and rhizomes. Plants are introduced into a constructed wetland basin by seeds, rhizomes, transplants or cuttings. Seeds may be introduced with borrowed topsoils, and will often be imported by wind, water and biological activity. The use of plant part propagules hastens establishment of the first year standing crop, but at some considerable expense. This method is necessary for SSF systems with barren media. Natural re-growth will occur in FWS systems, and will unavoidably bring diversity into the ecosystem. It has been repeatedly demonstrated that plants are necessary for optimal treatment efficiency, but there is no conclusive body of knowledge on the relative effectiveness of different species. The choice of species has thus far been very subjective. For example, Phragmites is the preferred plant in most of Europe, but is banned in many of the United States as a noxious weed. Hydrology and hydraulics The most important principle of design is that the water must be managed correctly for optimal treatment. That entails maintaining the desired operating depth and detention time, while preventing short-circuiting. But even the best designs are subject to atmospheric influences of rain, evapotranspiration, ice formation and snow melt. Depth control results from the correct combination of bottom contouring, layout geometry, media selection and an outlet structure design. Contouring has often proven to be problematic, especially for large wetlands, because of the need for close tolerances over large areas. Layouts favouring high length to width have been shown to have merit in treatment efficiency, but cannot compensate for the loss of area engendered by a sinuous path design. Further, high length to width increases internal head loss, and transfers depth control from the outlet structure to the vagaries of vegetation density or media hydraulic conductivity. Indeed, many of the first SSF wetlands

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Wetlands for Contaminant and Wastewater Treatment were built without regard for the hydraulic conductivity constraints implicit in the selection of the media. As a consequence, those early systems displayed either over-flooding or drying-out of portions of the wetland. The best designs provide for control of water depth by means of settings at the outlet structure, which may be selected to eliminate sensitivity to physical and biological variability internal to the wetland. It is obvious that optimal functioning requires that water has equal opportunity to contact all zones of the wetland. Channels that bypass the action zones of the ecosystem impair treatment. To a degree, utilising cells in series may ameliorate short-circuiting, because the collection and redistribution of the water intercepts preferential flow paths. However, a large body of information on the flow patterns in treatment wetlands indicates that there are unavoidable hydraulic inefficiencies inherent in wetlands of both types (Figure 20.6). Some parcels of water will follow fast paths; others follow slow paths (Wörman and Kronnäs 2005). The resultant distribution of

1.2 1.0

Lakeland Grand lake

0.8 0.6 0.4 0.2 0.0 0

1

2 3 Time, nominal detentions

4

Fig. 20.6 Residence time distributions for a FWS wetland (Lakeland, FL, USA) and a SSF wetland (Grand Lake, MN, USA). These model results are fitted to data, with R2 = 0.87 for Lakeland and R2 = 0.94 for Grand Lake. Grand Lake has an element of pure transport delay, representing the first tracer breakthrough at 5 days, compared to the cell detention of 14.8 days. Lakeland has a short 0.6 day transport delay, compared to the cell detention of 5.7 days.

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447

detention times shows that some water leaves after only a small fraction (c. 5–25%) of the average detention, while some is detained for up to three times the average. This universal observation of non-ideal wetland flow patterns is in direct conflict with the common sizing presumption of plug flow, and places severe and previously unrecognised constraints on wetland design procedures (Kadlec 2000a; Kadlec and Wallace 2008). Atmospheric processes are important modifiers of wetland processes. Evapotranspiration and rain have two effects: lengthening and shortening, respectively, of detention time, and concentration and dilution, respectively, of dissolved constituents. The use of an average flow rate in design compensates for altered detention time, but not for dilution or concentration. The fractional error due to flow averaging is approximately equal to fractional augmentation. Thus, if 25% of the inflow evaporates, use of an average flow predicts concentrations 25% lower than required by the mass balance. If rain adds 25% to the flow, use of an average flow predicts concentrations 25% higher. In cold climates, wetlands are subject to the influences of snow and ice. Snow provides a beneficial effect in the form of insulation that prevents freezing under all but the most extreme conditions in SSF wetlands. However, the spring snow melt adds considerably to the water flow for a brief period, during which contaminant reductions are altered. FWS wetlands typically experience ice formation, but again the presence of a snow cover limits the thickness to a fraction of that occurring on neighbouring lakes. Standing dead vegetation is effective in accumulating more snow than on barren surfaces. Nonetheless, winter storage may be required in extremely cold climates. Processes A large variety of wetland processes contribute to contaminant reduction in wetlands (Table 20.1). The several chapters of Section II discuss these in for wetlands in general. However, in the context of constructed treatment wetlands, it becomes

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Table 20.1 Proportioning of wetland pollutant processes. Areal processes (A) are associated with the areal extent of the wetland, and are related to soils, sediments, the water surface, plants and litter. Water column processes (W) are proportioned to the wetland water volume, and are related to immersed leaves and standing dead and suspended organisms. Process

TSS

BOD

TP

TN

Accretion Anaerobic digestion Carbon supply Chemical reaction Complexation Decomposition Deposition Diffusion Fixation Fungal processing Microbial and algal Benthic mat Epiphyton Metaphyton Plankton Rhizosphere Oxygen transfer Oxygenation Root Submerged Plant uptake Resuspension Sorption Volatilisation

A

A A

A

A

NOx-N

Metals A

A A

A A

A A A

A W A A A

W A A A

A W

A

A A A A A

A W A W A

A W A W A

A W A

A A A

important to understand how design influences the relative effectiveness of each process. At a first level of importance is understanding of the relative importance of wetland area and wetland water volume. For a fixed water flow and depth, more area provides more volume and hence a larger detention time. However, it is also possible to obtain greater detention by operating a deeper wetland water body. Some wetland processes are area-specific: they increase with wetland area but not with water depth, and are labelled ‘A’ in Table 20.1. An example of an areal process is plant uptake. Other processes are volume-specific: they increase with water volume, not with wetland surface area, and are labelled ‘W’ in Table 20.1. An example of a water column process is complexation of metals by dissolved humic substances. In design, if one wishes

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NH4-N

A A A

A A A A

A A

A

A W A W A A

A W A W

A W A

A W A A A

A

A

A A

to emphasise areal processes, then detention should be adjusted by adjusting wetland area. Adjusting depth fosters emphasis on water column processes. A second feature of treatment wetland process optimisation is the existence of supply constraints for some removal pathways. For instance, oxygen is necessary for nitrification; carbon is necessary for denitrification. Wetlands possess some measure of ability to meet these constraints. In the context of treatment wetland design, steps may be taken to ensure that such constraints are met. A third feature of treatment wetlands process design is the ability to acknowledge seasonality in meeting treatment objectives. The rates of some wetland processes, notably those that are biologically mediated, respond to season and temperature. Sometimes, there is a slowing at

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Wetlands for Contaminant and Wastewater Treatment colder temperatures, as for nitrogen reactions; sometimes season is more important, as for plant uptake during the spring growth surge. Design can compensate for such annual variability, either by providing enough area to meet bottleneck conditions, or by providing water storage to adjust flows to meet objectives. Winter water storage in lagoons is an example of the latter.

12

Data Cyclic model

10 8 6 4 2 0 0

1

2

3

4

5

P E R F OR M AN CE Spatial and temporal attributes The long detention times and relatively large areal extents associated with treatment wetlands lead to the need to consider spatial phenomena and their timing. Performance analysis is further complicated because deterministic phenomena are masked by stochastic events that cause data scatter. The economic importance of treatment wetlands has served to focus research on these facets of wetland behaviour. An often overlooked issue is the timing of wetland water quality sampling. A typical average transit time for water is days or weeks, depending on the specific design targets. Samples are often taken at the same time at inlet and outlet, and compared to compute removal percentages. Such a single point comparison is flawed by the fact that it is unlikely that the exit sample started its travel at the current inlet concentration, or that the current inlet sample will experience the same set of conditions as the outlet sample. Indeed, as pointed out in the preceding section, the current composite outlet water had its origins over 0.1–3.0 nominal detention times earlier. This synoptic error pervades the literature, and may be avoided by appropriate averaging, which is to average inlet and outlet concentrations over many detention times, and then calculate removals. Many wetland variables undergo cyclic seasonal behaviour, in response to the annual cycles in water flows and temperature, inlet concentrations, meteorology and solar radiation (Kadlec 1999a). An understanding of these cycles requires a good deal of information, preferably gathered over several consecutive years. Infrequent data

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6 7 8 Month

9

10 11 12 13

Fig. 20.7 Effluent CBOD5 from the Columbia, MO, USA treatment wetland. Over 3300 samples taken over 13 years are represented. Influent CBOD5 averaged 12.6 mg L−1; effluent averaged 4.6 mg L−1 (CBOD5 = 4.6(1 + 0.20 cos[0.0172(t − 147)], R2 = 0.16). (City of Columbia, unpublished data.)

0.40 0.35 0.30 0.25 0.20 0.15 0.10 0.05 0.00

−2

−1

0

1

2

3

4

5

CBOD5 deviation from seasonal norm (mg L-1) Fig. 20.8 Departure from the seasonal norm for CBOD5 at the Columbia, MO, USA wetlands, as represented in Figure 20.7. About one sixth of the variability in CBOD5 was accounted by the sinusoidal annual pattern; the remaining 84% is represented by this frequency distribution. Mean annual outlet CBOD5 = 4.7 mg L−1.

over one year are likely to be too strongly masked by stochastic variability. Such data may be detrended using annual cosine waves, thus separating the deterministic annual trend from the probabilistic components (Figures 20.7 and 20.8). The data sometimes contain counter-intuitive results, as is the case for the Columbia CBOD5 data, a 3-day detention system. Figure 20.7 shows a maximum outlet CBOD5 of about 6 mg L−1 at

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yearday 160, the beginning of June. The inlet CBOD5 for this wetland was essentially constant at 12.6 mg L−1 at approximately the same time, reflecting the seasonality of the pre-treatment. The minimum removal of 59% occurs in the middle of March, the maximum of 67% at the beginning of September. Wetland water temperature is apparently not the determinant, because it is at a minimum in January and maximum in July. Much of the variability in the data from Columbia is probabilistic – about 84%. The distribution has a lognormal appearance, with a fairly wide spread. Importantly, 10% of the data are more than 50% greater than the annual mean of 4.6 mg L−1. Spatial variability in treatment wetlands is also composed of deterministic and stochastic components. In addition, the (erroneous) historical concept of plug flow leads to the (erroneous) concept that distance and detention time are interchangeable under the transformation time is equal to distance/velocity. Various investigators have (1) measured concentrations on transects in the flow direction; (2) measured inlet and outlet concentrations for side-by-side wetlands of different lengths and the same velocity; (3) measured inlet and outlet concentrations for side-by-side wetlands of the same length and different velocities; (4) measured inlet and outlet concentrations for side-by-side wetlands of the same area and velocity but different depths; and (5) measured concentrations as a function of time in wetlands operated in a no-flow, batch mode. It is abundantly clear that these are not all equivalent experiments, as some of the wetland design literature recognises (Stein et al. 2003; Kadlec and Wallace 2008). However, all indicate two general features: a rapid decline in concentration with distance or time, followed by a levelling off to an asymptote at long distances or times. Regardless of the underlying wetland phenomena, such behaviour can be well represented by an exponential decline to the plateau value. This behaviour is illustrated in Figure 20.9 for total nitrogen, in systems investigated under protocols (1), (3) and (5). Total nitrogen tends to an asymptote of about 1.5 mg L−1,

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1.20 Ruakura, New Zealand Tres Rios, Arizona Humboldt, Saskatchewan

1.00 0.80 0.60 0.40 0.20 0.00 0

5

10 15 Time (days)

20

25

Fig. 20.9 Fractional decrease in total nitrogen with detention time in wetlands. The Tres Rios data are from points along the transect in a single FWS wetland, and reflect time of passage (City of Phoenix, unpublished data). The Ruakura data are from parallel SSF cells with different detention times (Tanner et al. 1995). The Humboldt data are from three replicate FWS wetlands operated in the batch mode, in which detention is elapsed time (Lakhsman 1981). The asymptote for Humboldt is 3.0 mg L−1.

in the form of organic nitrogen, as has been found for a large number of wetlands (Kadlec and Knight 1996). The fact that data from these three very different wetlands fall on approximately the same curve may be coincidental. Pollutant removal The central issue in treatment wetland design is pollutant reduction. There are two basic measures: concentration reduction, a flow-independent quantity, and load reduction, which involves the water quantities as well as the pollutant concentrations entering and leaving the wetland. Concentration reductions vary from substance to substance (Table 20.2), and are often dependent on inlet concentration and water detention time, or equivalent hydraulic loading rate. Removals rarely reach more than 90–95%, because of stochastic processes, wetland background concentrations and hydraulic imperfections.

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Table 20.2 Inlet and outlet concentrations (mg L−1) from constructed reed beds in several European countries and the USA. Number of beds shown in parentheses.

Austria (1) Czech Republic (2)

out in out

Denmark (3)

in out

Germany (1)

in out

Poland (1)

in out

UK (4)

in out

USA (5)

in out

BOD

COD

15 85 12 (45) 128 18 (69) 248 29 (39) 158 24 (7) 176 40 (32) 29 12 (23)

49 202 54 (44)

(63) 430 95 (107) 669 101 (5) 342 113 (21)

TSS

TN

TP

65 10 (42) 163 27 (66)

46.4 26.7 (31) 36.7 21.0 (66) 115 60 (9) 80 33.6 (7)

6.6 3.2 (32) 9.1 5.8 (52) 15.9 4.8 (21)

289 39 (5) 154 31 (34) 66 13 (23)

19.4 9.4 (12)

8.9 4.7 (12) 4.5 3.5 (8)

NH4-N

NOx-N

15.4 30.2 16.5 (32) 21.0 14.1 (46) 80.5 29.0 (45) 45 19.8 (4) 28.7 25.1 (34) 5.9 4.5 (18)

8.0 5.0 5.6 (20) 4.1 2.0 1.9 2.5 (39)

1.5 5.6 (20) 4.6 1.4 (12)

(1) Vymazal et al. (editors) 1998; (2) Vymazal (2002); (3) Schierup et al. (1990); (4) Cooper et al. (1996); (5) Knight et al. (1993).

Suspended solids Total suspended solids (TSS) contribute to wetland water quality both directly and indirectly, as a carrier of a number of incorporated and sorbed substances. Particles usually settle easily in the long travel times in the ecosystem, and become trapped in litter in the FWS wetlands, or in crevices in the SSF wetland media, together called filtration. Wetland processes produce particulate matter via decomposition of detritus from dead plants and algae and invertebrates. The formation of chemical precipitates such as metal sulphides adds to the internal production of solids. Wetland sediments are often easily disturbed, so that bioturbation by fish, mammals and birds can re-suspend these materials and lead to high TSS in the wetland effluent. Wetland particulate cycling is large and almost always overshadows additions. TSS is deposited near the inlet of both FWS and SSF treatment wetlands (Tanner and Sukias 1995).

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TSS background concentrations are rarely irreducible leftovers from incoming water; but are the result of the wetland processes. Most surface flow wetlands are sufficiently large to approach background levels. Typical FWS background concentrations are in the range of 3–10 mg L−1, but may be as low as 1–3 mg L−1 for SSF wetlands used for effluent polishing (Cooper et al. 1996; Kadlec and Wallace 2008). The rate at which background is approached is determined by the settling characteristics of the incoming and cycled solids. Under most circumstances, detention times that are adequate to reduce other contaminants will be sufficient to reduce TSS to background levels. The exceptions involve colloidal solids that remain in suspension for long periods. If the areal loading is less than 30 kg ha−1 d−1 of solids for a FWS wetland (USEPA 2000a), or 100 kg ha−1 d−1 for a SSF wetland (Wallace and Knight 2006), the exit concentration will generally be less than 30 mg L−1.

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Biochemical oxygen demand Treatment wetlands are quite effective at reducing biochemical oxygen demand (BOD), via aerobic and anaerobic processes. The organisms responsible for BOD consumption are associated with biofilms on plant, detritus and matrix surfaces, akin to other attached growth treatment processes, such as trickling filters. The oxygen required for aerobic degradation is supplied from the atmosphere, either directly by diffusion or indirectly from oxygen leakage from the macrophyte roots. Humic substances and other forms of carbon, which contribute to colour in the water, are characteristic of wetlands. The BOD test responds to these naturally occurring compounds, as well as to nitrogenous compounds. Carbonaceous biochemical oxygen demand (CBOD) is measured by inhibiting the oxidation of nitrogen species. The active wetland carbon cycle adds carbonaceous matter to the water column, as the by-product of detrital decomposition, creating background BOD in the wetland waters. The plateau concentrations in natural and treatment FWS wetlands are 3–10 mg L−1, but can be as low as 1–2 mg L−1 in SSF wetlands used for effluent polishing. Wetlands typically respond to increasing BOD loads by increasing the processing rates, but increased export of the contaminant still occurs. If the external areal loading is less than 40 kg ha−1 d−1 of BOD for a FWS wetland (USEPA 2000a), or 80 kg ha−1 d−1 for a SSF wetland (Wallace and Knight 2006), the exit concentration will generally be less than 30 mg L−1. Some perspective on the internal production of BOD is gained by considering the fate of the wetland biota after death and decomposition. An annual production of 5000 g m−2 of dry biomass is not uncommonly high for a treatment wetland, most of which ultimately undergoes decay. If all this biomass were to decay under water, the equivalent BOD internal loading would be much higher than typical external (added) BOD loadings. However, some decay occurs for the standing dead material while it is in air, and some is digested anaerobically. Nevertheless, internal BOD production and removal is high, and strongly influences

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background concentrations and apparent removal rates (Kadlec and Wallace 2008). Nitrogen Nitrogen removal processes interconnect the several species of nitrogen that are present in wetland waters: organic nitrogen (ORGN); oxidised nitrogen (ON), which is chiefly nitrate (NO3-N); and ammonia nitrogen (AN), which is chiefly ammonium (NH4-N). The total of all water-borne species is total nitrogen (TN). In addition, significant amounts of nitrogen are to be found in wetland soils, sediments and macrophytic vegetation. Also, wetland sediments have a significant sorption capacity for ammonium nitrogen. Mineralisation, or ammonification, is microbially mediated, and converts ORGN to AN. Nitrification is also microbially mediated, and converts AN to ON. Denitrification is also microbially mediated, and converts ON to dinitrogen gas, which is mostly released to the air above the wetland. These processes are individually and collectively sensitive to either or both season and temperature. Nitrification is subject to an oxygen supply requirement, and denitrification is subject to a carbon supply requirement. The rates of the processes typically decrease in the order: denitrification, which is greater than mineralisation, which is greater than nitrification. As a consequence of the serial nature of the transformations, and the rate differentials, the species composition of the wastewater to be treated is a critical factor in performance. Organic nitrogen must pass through the entire sequence, while nitrate is eliminated in the last and fastest step. In systems with a high proportion of incoming organic nitrogen, the ammonification step can out-compete the slower nitrification step, thus increasing the ammonium concentration during passage through the wetland. Some early treatment wetland literature speculated that the source for increase was from wetland sediments (USEPA 1993); however, subsequent studies have shown that waterborne organic nitrogen may also be a source (Burgoon et al. 1999).

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Wetlands for Contaminant and Wastewater Treatment Treatment wetland waters, like those of natural wetlands, contain a small background amount of organic nitrogen. Both SSF and FWS systems have background ORGN of about 1.0–1.5 mg L−1. If the external areal loading is less than 3.0 kg ha−1 d−1 of TN for a FWS or SSF wetland, the exit concentration will generally be less than 10 mg L−1 (Kadlec and Knight 1996). Phosphorus Phosphorus (P) may be sustainably removed at low rates in FWS treatment wetlands (Kadlec and Knight 1996; Kadlec and Wallace 2008). Higher rates may be obtained in SSF wetlands employing media with high P sorption capacity (Maehlum et al. 1995), but such systems have a finite lifetime before saturation of the media. The largest existing treatment wetlands are FWS systems targeting P removal from agricultural runoff (see Section VII, Chapter 42). These rely on the sustainable accretion of new soils and sediments containing refractory organic and calcium bound phosphorus. Calcitic materials are produced in hard water systems via precipitation of calcium minerals, and by co-precipitation with carbonates. These processes are mediated by the algal component of the biotic assemblage. Precipitates based on iron (Fe) and aluminium (Al) chemistry may form in waters which contain continuing supplies of these dissolved metals, but the stability of the solid Fe and Al compounds is contingent upon maintenance of high redox potential and high pH (Patrick and Khalid 1974). The organic component of the newly accreting solids is formed as residual from the decomposition of microorganisms and macrophytes (Reddy et al. 1999, 2005). Biogeochemical cycling of P through the microbial and algal compartments of the wetland is rapid, as are the inter-conversions among particulate, organic and soluble reactive P. Total accretion rates of P are typically in the range of 0.1–5 gm P m−2 a−1, and increase with increasing P concentrations in the water column. The accreted sediments typically contain 400–2000 mg P kg−1 dry solids.

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453

Both FWS and SSF wetlands may store phosphorus as a structural component of biomass and as sorbed P on antecedent soils. The capacity of these storages is limited by the maximum standing crop and sorption capacity respectively. A newly planted wetland, and some existing wetlands, show a growth increase response to new inputs of P and other nutrients, and remove P from the water to achieve new growth. Biomass expansion is usually complete within a year or two. The magnitude and direction of P transfer to wetland sorption sites is dependent on the type and antecedent condition of the wetland soils. Treatment wetlands built on previously agricultural soils may be loaded with phosphorus, and release some portion upon wetting (Kadlec and Knight 1996). Conversely, the media in SSF wetlands maybe selected to have a large P binding capacity. Sorption may therefore serve as an early premium for P removal, or as the design basis for Type C wetlands. Pathogens Human and animal wastewaters contain bacteria, parasites and viruses, which pose threats to human and wildlife health. Many of these pathogenic microorganisms will survive the various pre-treatment steps prior to a treatment wetland. The wetland environment is inimical to many of these enteric organisms. They attach to sediments, become subject to predation, are exposed to ultraviolet radiation, and may sometimes be subjected to damaging temperature regimes. The combination of these destructive processes leads to removal in the wetland, but there is also production by warm blooded wetland fauna. As a consequence, FWS wetlands possess relatively small numbers of pathogens, and removal is not complete. SSF systems place the water underground, and minimise the potential for contact by humans or animals. This feature is preferred for treatment of settled sewage and other wastewaters that have a minimum of pre-treatment for pathogens. There is little or no internal production in SSF wetlands, and removals can proceed to quite low residual levels.

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Metals Some common metals are essential to the functions of natural wetland ecosystems, notably calcium, magnesium, iron and manganese. Calcium is a vital ingredient in algal processes, and can form the basis for sediment inorganic chemistry in hard water wetlands, in which the soils are calcitic muds. Iron is important in many other wetlands, as an integral part of the sediment and soil chemistry. Other heavy metals ordinarily play a minor role in the functioning of natural wetlands, but interact strongly with the sediments and soils (Sobolewski 1999). Divalent cations may be ion exchanged out of water solution, onto soil proton donor sites. The lower soil horizons frequently support sulphate reduction to sulphide, and provide the environment for metal sulphide precipitation. As a consequence, wetlands have the ability to trap heavy metals and store them in soils. The capacity can be large, and if metal loadings are large, high metal concentrations can be created in the surface soils and sediments. Therefore, potential hazards can be created for wildlife health if high metal concentrations are passed into treatment wetlands. Careful evaluation of those risks is a necessary part of treatment wetland design for heavy metal trapping.

M ODE L L IN G AN D DE S IGN Descriptive techniques Wetland models of greatly varying complexity have been set forth over the past three decades. The more complex models include dynamic behaviour of the various ecosystem compartments and processes (Howell et al. 2005), but these require very large amounts of data for proper calibration. Calibrated compartmental models will presumably provide more details of internal allocations of chemicals, but such detailed deterministic models will not necessarily provide more accurate descriptions of overall wetland performance,

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because of stochastic factors. At this point in the evolution of treatment wetland technology, only simple models can be calibrated for most operational systems. To forecast water quality performance of treatment wetlands, two principal features must be jointly considered: hydraulics and pollutant removal. Previous literature utilises concepts of Darcian flow in SSF wetlands (Fisher 1990), and of vegetated open channel flow in FWS wetlands (Hosokawa and Horie 1992). Previous literature suggests first order, irreversible pollutant reduction removal models for treatment wetlands, which may be either area-specific, and thus determine the necessary wetland acreage (Vymazal et al. 1998), or volume-specific, and thus determine the wetland water volume (USEPA 1993). These equations purport to represent the wetland output concentrations in response to inlet concentrations, flow rate and area or volume. However, wetland performance also includes a great deal of variability that is not predicted by the average values of these forcing variables. Unpredictable events, such as the fluctuations in input flows and concentrations, changes in internal storages, as well as weather and animal activity contribute to observed performance. In general, wetland state variables, such as concentrations, may be represented by a seasonally varying periodic component plus a stochastic component (Kadlec 1999a): C = Cavg[1 + A · cos(w(t − tmax))]

(20.1)

where: w annual frequency = 2π/365 = 0.0172 day−1 A amplitude of the annual cycle, as a fraction of the annual average C instantaneous value of the variable Cavg annual average value of the variable t year day, from 0 to 365 tmax year day for maximum variable value E additive stochastic component of the variable Figures 20.7 and 20.8 show an example of data, and the representation contained in Equation (20.1).

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Wetlands for Contaminant and Wastewater Treatment Design often contemplates a stable period of operation, over which input and outputs are averaged. If enough data points are included within the season of the averaging, then the stochastic component is eliminated. If the averaging period is one or more calendar years, then the seasonality is eliminated, regardless of the mechanisms involved. A frequent observation is the presence of a decreasing concentration along the flow direction inside the treatment wetland or, correspondingly, a decrease with increasing detention time (Figure 20.9). The rate of decrease often decreases with distance or time and a plateau is approached at long distances or times, if the wetland is large enough. These observations are consistent with a first order model, which contains a rate proportional to concentration. They are also consistent with one of several hydraulic models, ranging from plug flow to a few tanks in series. The plug flow assumption is that elements of entering water all pass through the wetland at precisely the average detention time. However, many independent wetland tracer studies have produced a bell-shaped distribution with a long tail, which is not characteristic of plug flow, but rather a gamma distribution of travel times (Figure 20.6): f(t) = a(lt)β exp(−lt)

(20.2)

where: f travel time probability density, or residence time distribution function, 1/d t travel time, d a, b, l constants t average detention time, b/l Given this distribution of detention times, the pollutant reduction for a first order decay to background may be written as (Kadlec and Wallace 2008):

Co − C * = Ci − C *

 kt  1 +  hb  

where: Co outlet concentration, mg L−1 Ci inlet concentration, mg L−1

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−b

(20.3)

1000 100 10 120

1 0.1 0.01 1

10

100

1000

10 000 100 000

TKN load (gN m−2 yr−1) Fig. 20.10 Ammonium nitrogen outlet concentrations as a function of areal TKN loading for 168 FWS wetlands. The data is partitioned by inlet concentration ranges in mg L−1. Each point represents the entire performance history for a single wetland. (Data source: Kadlec and Wallace 2008.)

C* background concentration, mg L−1 h wetland free water depth, m k areal removal rate constant, m d−1 q hydraulic loading rate, m d−1 t detention time, days A second approach for relating effluent concentrations to inlet flows and concentrations is a graphical display of concentration versus loading (USEPA 2000a; Wallace and Knight 2006; Kadlec and Wallace 2008). The inlet loading is defined as the rate of pollutant addition per unit wetland area. It is necessary to also identify the inlet concentration as a primary variable, leading to data clusters on Figure 20.10. A monotonic increase is observed for both intra- and inter-wetland data sets (see Figure 20.10 for example). It may be shown that Equation (20.3), for various inlet concentrations and hydraulic loadings, will produce predictions that cluster around such an increasing trend (Kadlec 1999b). This trend means that higher inlet concentrations, as well as higher water flows, produce higher outlet concentrations. Trend graphs such as Figure 20.10 are useful for visualising the performance of many treatment wetlands, and identify the uncertainties inherent in design. New designs that forecast points above the central

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160 8/25/93 9/2/93 8/25/93 Model 9/2/93 Model

140

BOD (mg L–1)

120 100 80 60 40 20 0 0.0

0.2

0.4 0.6 Fractional distance

0.8

1.0

Fig. 20.11 Fit of the first order model to BOD data from American Crystal Sugar’s treatment wetlands in Hillsboro, ND, USA. Circles and solid line are calibration data and fit; triangles and dashed line are validation data and fit. Note that one internal point is poorly predicted. Data from Anderson (1996). There are three tanks in series, k = 102 m a−1 and C* = 3 mg L−1.

trend line would be conservative, in that most operating systems loaded at similar rates would have lower effluent concentrations. Conversely, designs located below the central tendency would be accompanied by more risk.

Equation (20.3) presumes that such storages are at a steady, equilibrium value, which is not true on a short time scale. Long-term averaging is again required to overcome this difficulty. Many previous pollutant models for treatment wetlands have adopted the plug flow assumption, which leads to an exponential decay of concentration to wetland background. This corresponds to b = ∞ in Equation (20.3), whereas wetland data frequently indicate 2 ≤ b ≤ 4. The rate constant is very sensitive to the value of b. For instance, the plug flow k = 66 m a−1 for the data in Figure 20.11, as compared to the value k = 102 m a−1 for b = 3. The effect of this sensitivity is to impair the transferability of rate constants from one situation to another within the same wetland. A second effect is a strong dependence of the rate constant on hydraulic loading rate (Kadlec 2000a). In design, Equation (20.3) is used to compute the detention time, or equivalent hydraulic loading, required to lower the inlet concentration by the desired amount. This calculation does not include stochastic effects, which must be accommodated by adjustment of the design target concentration. The value of the first order rate constant is temperature sensitive in some instances, notable for nitrogen transformations, and thus design must focus on the seasonal ‘bottleneck’ in performance.

Calibration and variability Data fitting to Equation (20.3) is normally quite good for a specific data set, but stochastic variability often prevents accurate predictions in the design mode (Figure 20.11). Care must be exercised in selection of calibration data sets. Synoptic error occurs when calibration data are acquired at essentially the same time for inflows and outflows, because nominal detention times in wetlands are often in the order of several days. Inputs and outputs cannot simply be offset by one detention time, because of the blurring implicit in Equation (20.2). The remedy is to average the calibration data over several detention times. Temporary storage error occurs when growth, release or sorption undergo temporary excursions due to season, temperature or other factors.

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Constraints Both the deterministic and stochastic components of wetland pollution reduction are contingent on chemical, physical and biological constraints (Table 20.3). Equation (20.3), or any equivalent, presumes that the necessary constraints are met. For instance, the removal of oxidised nitrogen via denitrification requires both a population of denitrifying bacteria and a carbon source to feed them. If both are present, then a concentration and flow-based formula may be calibrated. Denitrifying bacteria pervade almost all soils and sediments, and populations develop with great rapidity in response to the chemical environment. But if the carbon source is absent, the removal will be zero. This particular example

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Table 20.3 Constraints on pollutant removals in treatment wetlands for individual processes. More than one process may be involved in the reduction of a pollutant. Constraint

Type

Magnitude

Oxygen for BOD reduction Oxygen for nitrification Alkalinity for nitrification Accretion for phosphorus burial Carbon for denitrification Sulphide for metal binding Nitrogen for plant growth Phosphorus for plant growth Sorption capacity for phosphorus Cation exchange capacity for metal binding Existence of microbial communities

Sustained Sustained Sustained Sustained Sustained Sustained Sustained Sustained Life limiting Life limiting Sustained

1.0 mg L−1 per mg L−1 BOD 4.7 mg L−1 per mg L−1 NH4-N 7 mg L−1 per mg L−1 NH4-N c. 0.2% of dry sediment 1.1 mg L−1 per mg L−1 NOx-N 16 gm per equivalent c. 2% of dry matter c. 0.2% of dry matter Substrate specific Substrate specific Chemical specific

has happened in more than one treatment wetland (Gersberg et al. 1983). An adequate oxygen supply is perhaps the most frequently unmet constraint. The large oxygen demand exerted by ammonia and BOD in strong wastewaters often cannot be met by some types of treatment wetlands, notably horizontal SSF systems. This fact has caused the development of vertical and horizontal systems operated intermittently to bring air into the root zone (Cooper 1999; Behrends et al. 2001).

Implementation Alternatives analysis Evaluation of options for treatment of a contaminated water flow proceeds at two levels: comparison of wetlands with other means of treatment, and comparison of the various wetland alternatives. At the first level, wetlands are in competition with, or may be added to, other technologies, such as activated sludge systems, lagoons and land application. The primary considerations in this phase of evaluation are land availability and capital and operating costs. Wetlands require more land than most conventional alternatives, and therefore are not feasible in crowded circumstances. The capital cost of the land is an important factor, particularly for

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large systems. But operations are simpler and less costly than for mechanical plants. Life cycle energy requirements are much lower for Type A treatment wetlands (Brix 1999). At the second level, there are several types of treatment wetlands that may be employed (Figure 20.1). SSF wetlands are preferred for wastewaters that pose a health contact threat to humans or wildlife, but they are much more expensive than FWS systems for the same degree of treatment. FWS systems provide more ancillary benefits, such as wildlife and human use. If treatment is constrained by oxygen supply, then the vertical flow bed is a good candidate for inclusion. In this phase of consideration, multiple wetland types and combinations of wetlands with other natural systems are evaluated. A common outcome of this phase is the choice of a combination of vertical and horizontal flow wetlands, or a combination of lagoons and wetlands, or a combination of mechanical treatment with wetland polishing. Regulation The development of treatment wetland technology has placed stress on regulatory agencies. In addition to the growing pains associated with any new technology, there are perceived issues relating to wetland preservation, restoration

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and mitigation. A constructed wetland treating strong effluent is typically regarded as a treatment system, and regulated as such. But wetlands created to polish high quality wastewaters have most or all of the attributes associated with mitigation wetlands, including biological diversity, habitat and human use. Guidelines for establishing such multi-purpose projects have been set forth (USEPA 2000b). There is disparity among the types of regulation of the various types of treatment applications. Large volume point source wastewaters are subject to the permitting requirements common to all alternatives for treatment, such as the National Pollutant Discharge Elimination System (NPDES) in the USA. But stormwater treatment wetlands are frequently regulated via technology based standards, which means they are built to a design that implies the level of performance. Wetlands treating small flows, from individual or small clusters of residences or businesses, typically discharge to subsurface aquifers. These are viewed as enhancements of the treatment provided by septic tanks and infiltration fields, and are regulated on the infiltration area and thickness of the vadose zone. Treatment wetland performance is often evaluated by comparison to regulatory standards that include features affected by both the mean system performance and the stochastic variability. A typical operating permit or license will specify monthly and weekly means that are not to be exceeded beyond a specified frequency, and these limits may be different for different seasons of the year. For example, it might be required that the wetland produce weekly average outlet total phosphorus of less than 50 µg L−1 50% of the time. In order to account for this design requirement, information about the stochastic variability of the wetland performance must be considered – in other words, the ‘E’ portion of Equation (20.1) must be quantified. This is commonly accomplished by considering the frequency distribution of outlet concentrations, such as that shown in Figure 20.12. The cumulative distributions are often the ‘S’ shaped curves characteristic of lognormal behaviour. Examination of Figure 20.12 indicates that the system in question produces

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1.20 1.00 0.80 0.60 0.40

Inlet Outlet

0.20 0.00 0.000

0.050

0.100

0.150

0.200

0.250

Total phosphorus (mg L−1)

Fig. 20.12 Frequency distributions for wetland inlet and outlet total phosphorus concentrations for the ENR project in south Florida. Over 450 samples taken over 4 years are represented. (South Florida Water Management District, unpublished data.)

less than 50 µg L−1 TP 95% of the time, and therefore meets the example requirement easily. In design, frequency distributions from operating systems are used to define the required removal under average operating conditions. Continuing the same phosphorus example, if the target were to be 50 µg L−1 TP 95% of the time, it would be necessary to achieve 20 µg L−1 TP on average, or 50% of the time (Figure 20.12). Many other types of regulatory specifications are in common usage, and these all normally require knowledge of the performance distributions. Construction and start-up Wetlands are conceptually easy to build: excavate a shallow basin, add soil or media and plants, and bring in the water. In practice, each of those steps has been problematic. Earth moving and berm construction are routine, but extra care is required to establish the exceedingly level basin bottoms, with the very minute slopes that are needed. Placement of liners is also routine, but care is mandatory for media and soil placement, to avoid clogging dirt and excessive compaction. Ecosystem establishment begins with vegetation establishment, and is complete when all

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Wetlands for Contaminant and Wastewater Treatment components of the ecosystem are present and attendant biogeochemical cycles are in operation. Plant propagules may be purchased and added to the basin, or natural re-growth may be allowed to proceed, from incidental import and existing soil seedbanks. Transplanting of whole plants or plant parts is costly, especially for high-density insertion. Seeding is difficult, due to issues of source material availability, germination conditions, sowing procedures and time for development. Natural re-growth will occur in all circumstances, whether it is the sole process chosen, or occurs as natural invasion of planted systems. The ‘greening’ of a treatment wetland is only the first phase of development. Especially in FWS wetlands, the development of the standing dead and litter compartments is essential to performance, because these materials are substrate, refugia and food for the microbial communities that are vital to treatment. In some instances, treatment wetlands have been mulched with straw or other detritus, to hasten the establishment of the ‘brown’ portion of the ecosystem. This phase of development is not as critical for SSF wetlands, which possess large submerged surface areas associated with the media. These establishment processes are reflected in the evolution of treatment performance. A wetland with few plants and no detritus does not support a full complement of mechanisms, but operates in a storage mode, with nutrient uptake required to build the biomass to its operational level. Start-up is a term that has been used to identify the transient period of development during which treatment does not meet design objectives. Stabilisation is a term that has been used to describe the period after start-up during which treatment objectives are met, but performance trends continue. Start-up and stabilisation may be complete in a few months for SSF wetlands in warm climates that do not rely upon the macrophytic portion of the biogeochemical cycle, but can be as long as 5 years in cold climates for FWS systems utilising natural wetlands (Kadlec 1997).

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Economics Treatment wetlands are economically attractive in many circumstances, primarily because of low operating costs. However, the cost of building the system is often comparable to that for other treatment alternatives. Five categories of expense dominate the capital cost of treatment wetlands: land, earth moving, conveyance, liners and media. The first three categories are common to all types of treatment wetlands. Conveyance equipment costs are minimal if gravity flow can be utilised, but rise dramatically if large pumps are required, perhaps at both inlet and outlet. Only the SSF version incurs cost for gravel or other substrate, which is often a large portion of the cost. Liners are not always required. Engineering costs are likely to be higher than for conventional alternatives, in both money and time, because of extra steps in gaining permissions. The capital costs for both FWS and SSF wetlands in the USA have been found to depend on size (Kadlec and Wallace 2008). FWS: C = 194A0.690 R2 = 0.79 0.03 < A < 10 000 HSSF: C = 652A0.704 R2 = 0.75 0.005 < A < 20 (20.4) where: A wetland area, ha C capital cost 2006 in thousand of dollars. Capital costs for SSF systems are about three times higher than for comparable wetland area. For any particular contaminant, the required area may be higher or lower for SSF systems, but there is always a cost disadvantage to SSF systems. The capital cost to treat a specified flow is roughly the same for FWS and SSF wetlands. There are strong variations in costs from country to country, For instance, SSF systems in Portugal cost six times as much as in Central America. Operating costs for treatment wetlands are often much lower than for competing technologies. They require no chemicals, and Type A, FWS wetlands do not wear out, thus eliminating replacement costs. Cost categories include

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physical maintenance, flow and quality monitoring, and biological surveillance. Physical maintenance is minimal, usually being comprised of berm and structure maintenance, and animal control. It has been found that SSF wetlands must be refurbished periodically to alleviate clogging problems. The associated costs add substantially to the operating budgets (Cooper et al. 2008). Flow and quality monitoring, for purposes of compliance determination, is usually not burdensome. However, in some cases in North America, extraordinarily large ecosystem monitoring has been imposed on treatment wetlands, in connection with ecosystem issues rather than treatment issues. The range of treatment-related operating and maintenance costs is from a few hundred to a few thousand US$ per hectare per year.

S U M M AR Y Treatment wetlands of many diverse forms and functions have been found effective in providing water pollution control. Variants include surface and SSF, and many choices of substrate and vegetation. Biogeochemical cycling and accretion, microbial transformations and soil–water interactions all contribute to the ability of a wetland to improve water quality. Long detention times aid in removal, at the expense of large land areas for implementation. These natural systems are exposed to, and strongly interact with, meteorological and ecological processes. Domestic wastewater treatment was the predominant early use, and this application continues to grow in many countries. Newer applications include stormwater, animal waste, pulp and paper, petrochemical and food production; as well as leachate and mine drainage control, and groundwater cleanup. Operating costs for this land-intensive alternative are minimal, and capital costs are competitive. In some instances, ancillary habitat and human use benefits accrue to this treatment choice. The quantitative science and engineering basis for understanding and designing treatment wetlands currently

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rests on a very large body of information from operating systems, and continues to grow. Continued growth of the technology is assured in many regions of the world.

R EFER EN CES Anderson P.A. 1996. Advanced treatment of sugar beet process wastewater using constructed wetlands. In: Proceedings of 5th International Conference on Wetland Systems for Water Pollution Control. Institute for Water Provision, Universität für Bodenkultur Wein, Vienna, Austria, pp. X/2-1–X/2-8. Behrends L., Houke L., Bailey E., Jansen P. and Brown D. 2001. Reciprocating constructed wetlands for treating industrial, municipal and agricultural wastewater. Water Science and Technology 44(11–12), 399–406. Brix H. 1999. How ‘green’ are aquaculture, constructed wetlands and conventional wastewater treatment systems? Water Science and Technology 40(3), 45–50. Brix H., Arias C.A. and del Bubba M. 2001. Media selection for sustainable phosphorus removal in subsurface flow constructed wetlands. Water Science and Technology 44(11–12), 47–54. Burgoon P.S., Kadlec R.H. and Henderson M. 1999. Treatment of potato processing wastewater with engineered natural systems. Water Science and Technology 40, 211–216. Carleton J.N., Grizzard T.J., Godrej A.N. and Post H.E. 2001. Factors affecting the performance of stormwater treatment wetlands. Water Research 35(6), 1552–1562. CH2M Hill and Payne Engineering. 1997. Constructed Wetlands for Livestock Wastewater Management. Literature review, database and research synthesis. Gulf of Mexico Program, Stennis Space Center, Mississippi, USA. Cooper P.F. 1999. A review of the design and performance of vertical-flow and hybrid reed bed systems. Water Science and Technology 40(3), 1–10. Cooper P.F. and Findlater B.C. (editors) 1990. Constructed Wetlands in Water Pollution Control. Pergamon Press, Oxford. Cooper P.F., Job G.D., Green M.B. and Shutes R.B.E. 1996. Reed Beds and Constructed Wetlands for Wastewater Treatment. WRc, Swindon, Wiltshire. Cooper D.J., Griffin P. and Cooper P.F. 2008. Factors affecting the longevity of subsurface horizontal

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Wetlands for Contaminant and Wastewater Treatment flow systems operating as tertiary treatment for sewage effluent. In: Vymazal J. (editor) Wastewater Treatment, Plant Dynamics and Management in Constructed and Natural Wetlands. Springer, New York, pp. 191–198. Ewel K.C. and Odum H.T. 1984. Cypress Wetlands. University of Florida Press, Gainesville, FL. Fetter C.W., Sloey W.E. and Spangler F.L. 1978. Use of a natural marsh for wastewater polishing. Journal WPCF 50, 290–307. Fisher P.J. 1990. Hydraulic characteristics of constructed wetlands at Richmond, NSW, Australia. In: Cooper P.F. and Findlater B.C. (editors), Constructed Wetlands in Water Pollution Control. Pergamon Press, Oxford, pp. 21–32. Gersberg R.M., Elkins B.V. and Goldman C.R. 1983. Nitrogen removal in artificial wetlands. Water Research 17, 1009–1014. Hammer D.A. (editor) 1989. Constructed Wetlands for Wastewater Treatment. Lewis Publishers, Boca Raton, FL. Hokosawa Y. and Horie T. 1992. Flow and particulate nutrient removal by wetland with emergent macrophyte. Science of the Total Environment 126(Suppl.), 1271–1282. Howell C.J., Crohn D.M. and Omary M. 2005. Simulating nutrient cycling and removal through treatment wetlands in arid/semiarid environments. Ecological Engineering 25(1), 25–40. IWA Macrophyte Group. 2000. International Water Association working group on the use of macrophytes in water pollution control. Constructed Wetlands for Pollution Control: Processes, Performance, Design and Operation. International Water Association, London. James B.B. and Bogaert B. 1989. Wastewater treatment/ disposal in a combined marsh and forest system provides for wildlife habitat and recreational use. In: Hammer D.A. (editor), Constructed Wetlands for Wastewater Treatment. Lewis Publishers, Boca Raton, FL, pp. 597–605. Kadlec R.H. 1997. An autobiotic wetland phosphorus model. Ecological Engineering 8, 145–172. Kadlec R.H. 1999a. Chemical, physical and biological cycles in treatment wetlands. Water Science and Technology 40(3), 37–44. Kadlec R.H. 1999b. The limits of phosphorus removal in wetlands. Wetland Ecology and Management 7, 165–175. Kadlec R.H. 2000a The inadequacy of first order treatment wetland models. Ecological Engineering 15, 91–104.

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Kadlec R.H. 2000b. Using wetlands to diminish the effects of pollution. In: Water 2000, Proceedings of NZWWA Biennial Conference, Auckland, NZ, March 2000. Kadlec R.H. 2003. Integrated natural systems for landfill leachate treatment. In: Vymazal J. (editor), Wetlands – Nutrients, Metals and Mass Cycling. Backhuys Publishers, Leiden, pp. 1–33. Kadlec R.H. and Bevis F.B. 1990. Wetlands and wastewater: Kinross, Michigan. Wetlands Journal of Society Wetland Scientists 10, 77–92. Kadlec R.H. and Knight R.L. 1996. Treatment Wetlands. CRC Press, Boca Raton, FL. Kadlec R.H. and Wallace S.D. 2008. Treatment Wetlands (2nd edition). CRC Press, Boca Raton, FL. Knight R.L., Ruble R.W., Kadlec R.H. and Reed S.C. 1993. Database: North American Wetlands for Water Quality Treatment. Phase II Report prepared for USEPA, including diskette. Knight R.L., Kadlec R.H. and Ohlendorf H.M. 1999. The use of treatment wetlands for petroleum industry effluents. Environmental Science and Technology 33, 973–980. Lakhsman G. 1981. A Demonstration Project at Humboldt to Provide Tertiary Treatment to the Municipal Effluent Using Aquatic Plants. SRC Publication No. E-820-4-E-81, Saskatchewan Research Council, Saskatoon, Saskatchewan, Canada. Lampe L.K. and Grizzard T. 1999. Evaluating the Use of Constructed Wetlands in Urban Areas. Water Environment Research Foundation, Alexandria, VA. Litchfield D.K. 1993. Constructed wetlands for wastewater treatment at Amoco Oil Company’s Mandan, North Dakota refinery. In: Moshiri G.A. (editor), Constructed Wetlands for Water Quality Improvement. Lewis Publishers, Boca Raton, FL, pp. 485–488. Maehlum T., Jenssen P.D. and Warner W.S. 1995. Cold climate constructed wetlands. Water Science and Technology 32(3), 95–101. Mander U. and Jenssen P. (editors) 2003a. Constructed Wetlands for Wastewater Treatment in Cold Climates. WIT Press, Southampton. Mander U. and Jenssen P. (editors) 2003b. Natural Wetlands for Wastewater Treatment in Cold Climates. WIT Press, Southampton. Mander U., Vohla C. and Poom A. (editors) 2003. Constructed and Riverine Wetlands for Optimal Control of Wastewater at Catchment Scale. Tartu University Press, Tartu, Estonia.

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Means J.L. and Hinchee R.E. (editors) 2000. Wetlands and Remediation. Battelle Press, Columbus, OH. Mitsch W.J. and Gosselink J.G. 2000. Wetlands (3rd edition), Van Nostrand Reinhold, New York. Moshiri G.A. (editor) 1993. Constructed Wetlands for Water Quality Improvement. Lewis Publishers, Boca Raton, FL. Mudroch A. and Capobianco J.A. 1979. Effects of treated effluent on a natural marsh. Journal WPCF 51, 2243–2256. Mulamoottil G., McBean E.A. and Rovers F. (editors) 1998. Constructed Wetlands for the Treatment of Landfill Leachates. Lewis Publishers, Boca Raton, FL. NCASI (National Council for Air and Stream Improvement, Inc.) 2004. Use of Constructed Wetland Effluent Treatment Systems in the Pulp and Paper Industry. Technical Bulletin No. 876, NCASI, Research Triangle Park, NC. Nehring K.W. and Brauning S.E. (editors) 2002. Wetlands and Remediation II. Battelle Press, Columbus, OH. Nivala J.A., Hoos M.B., Cross C.S., Wallace S.D. and Parkin G.F. 2007. Treatment of landfill leachate using an aerated, horizontal subsurface-flow constructed wetland. Science of the Total Environment 380(1–3), 19–27. Patrick W.H. Jr. and Khalid R.A. 1974. Phosphate release and sorption by soils and sediments: effect of aerobic and anaerobic conditions. Science 186, 53–55. Reddy K.R. and D’Angelo E.M. 1994. Soil processes regulating water quality in wetlands. In: Mitsch W.J. (editor), Global Wetlands: Old World and New. Elsevier, Amsterdam, pp. 309–324. Reddy K.R., Kadlec R.H., Flaig E. and Gale P.M. 1999. Phosphorus assimilation in streams and wetlands: a review. Critical Reviews in Environmental Science and Technology 29, 83–146. Reddy K.R. and Smith W.H. (editors) 1987. Aquatic Plants for Water Treatment and Resource Recovery. Magnolia Publishing, Orlando, FL. Reddy K.R., Wetzel R.G. and Kadlec R.H. 2005. Biogeochemistry of phosphorus in wetlands. In: Sims J.T. and Sharpley A.N. (editors), Phosphorus: Agriculture and the Environment. Soil Science Society of America, Madison, WI, pp. 263–316. Schierup H.-H., Brix H. and Lorenzen B. 1990. Spildevandsrensning i Rodzoneanlæg. Botanical Institute, Aarhus University, Denmark, 87 pp.

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Schueler T.R. 1992. Design of Stormwater Wetland Systems. Metropolitan Washington Council of Governments, Washington, DC. Seidel K. 1966. Purification of water by higher plants. Naturwissenschaften 53, 289–298. Sobolewski A. 1999. A review of processes responsible for metal removal in wetlands treating contaminated mine drainage. International Journal of Phytoremediation 1, 19–51. Spangler F.L., Sloey W.E. and Fetter C.W. 1976. Experimental use of emergent vegetation for the biological treatment of municipal wastewater in Wisconsin. In: Tourbier J., Pierson R.W. Jr. (editors), Biological Control of Water Pollution. University of Pennsylvania Press, Philadelphia, PA, pp. 161–171. Stein O.R., Hook P.B., Biederman J.A., Allen W.C. and Borden D.J. 2003. Does batch operation enhance oxidation in subsurface constructed wetlands? Water Science and Technology 48(5), 149–156. Tanner C.C., Clayton J.S. and Updell M.P. 1995. Effect of loading rate and planting on treatment of dairy farm wastewaters in constructed wetlands. II. Removal of nitrogen and phosphorus. Water Research 291, 27–34. Tanner C.C. and Sukias J.P. 1995. Accumulation of organic solids in gravel-bed constructed wetlands. Water Science and Technology 32(3), 229–239. Tilton D.L. and Kadlec R.H. 1979. The utilisation of fresh-water wetland for nutrient removal from secondarily treated waste water effluent. Journal of Environmental Quality 8, 328–334. USDA (US Department of Agriculture) 1991. Nutrient Sediment Control System. Environmental Quality Technical Note N4, Soil Conservation Service, Chester, PA, USA. USEPA (US Environmental Protection Agency) 1993. Subsurface Flow Constructed Wetlands for Wastewater Treatment – A Technology Assessment. EPA 832-R-93-001 USEPA Office of Water, Washington, DC, USA. USEPA (US Environmental Protection Agency) 2000a. Constructed Wetlands Treatment of Municipal Wastewaters. EPA 625-R-99-010, USEPA Office of Water, Washington, DC, USA. USEPA (US Environmental Protection Agency) 2000b. Guiding Principles for Constructed Treatment Wetlands: Providing for Water Quality and Wildlife Habitat. EPA 843-B-00-003, USEPA Office of Wetlands, Oceans and Watersheds, Washington, DC, USA.

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Section IV Wetland Assessment: How can we Measure that Wetlands are Working?

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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21

Introduction – Methodologies for Wetland Assessment JOS EPH S. LARSO N

Environmental Institute, University of Massachusetts, Amherst, USA

IN T R O D U CT ION Human welfare depends on ecosystem goods such as food, timber, genetic resources and medicines. Research has demonstrated that ecosystems also provide services essential to human well being, such as water purification, flood control, coastline stabilisation, carbon sequestration, waste treatment, biodiversity conservation, soil generation, disease regulation, maintenance of air quality and aesthetic and cultural benefits. The importance of the need to assess the functions of these goods and services has been noted repeatedly in the more recent Proceedings of the Conferences of the Contracting Parties of the International Convention on Wetlands (Ramsar Bureau 2005). The scientific committee that explored the merits of launching a Millennium Assessment of the World’s Ecosystems identified a need for a system of international ecosystems assessment. They concluded that too little is known of the current state and future prospects of ecosystem goods and services, and that there is need for a system of international ecosystem assessment to support an integrated, predictive and adaptive approach to ecosystem management (Ayensu et al. 1999). Wetlands throughout the world provide most, if not all, of these goods and services (Hollis et al.

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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1988; Ramsar Convention 1990). This astounding mix has prompted wetland scientists over the last 25 years to develop methods and models for detecting, quantifying, or assigning relative rankings to the goods and services provided by wetlands. It should be noted, however, that wetland assessment has generated some controversy. Critics are concerned that formal assessment methods will be used to identify a percentage of the wetland resource that can be destroyed. Supporters of assessment view the process as a means of helping governments and NGOs make well-informed decisions about wetland protection and management. In those instances where the government has established basic wetland protection standards applicable to all wetlands, assessment can be a means to identify those wetlands that merit additional protection by identifying the presence of specific functions. In this section, wetland assessment is described comprehensively in terms of its historical development and practical applications, specifically in relation to legislative change (Smith, Chapter 24), and the problems and methods of assessing wetlands of many different scales and types according to scales that must have common meaning. Brinson (Chapter 22) and Maltby and Barker (Chapter 23) describe methods, designed for use in the USA and Europe respectively, based on hydrogeomorphic characteristics of the wetland landscape. The specific challenges facing wetland assessment in developing countries are examined by Roggeri (Chapter 25), who gives examples of how functions influencing the social,

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environmental, and economic objectives of such areas can be evaluated successfully. Underlying many assessments is the requirement to relate physical processes to economic values, often as measures of the goods and services that wetlands commonly provide for human societies. Central to this is the need to express the benefits of wetlands in a standardised and repeatable way using common terminology. Turner et al. (Chapter 26) describe approaches to this complex area, for example the Driving forces-Pressure-State-Impact-Response (DPSIR) approach, which is based on the functional diversity of the wetland, and requires the blending of economic, hydrological and ecological models. Economic valuation also must incorporate factors such as social welfare and sustainable economic development as key objectives in any comprehensive appraisal.

O R IG IN S O F AS S E S S M E N T Wetland assessment began when the first decisions were made to drain wet areas in order to convert them to some other use. Assessment procedures that address the wide array of intrinsic functions of wetlands are of recent origin. Following ratification of the Migratory Bird Treaty in 1916 between the United States and Great Britain, on behalf of Canada, wildlife biologists in Canada and the United States began to identify, catalogue and assess the wetlands in North America that are critical for migratory waterfowl nesting, migratory and wintering habitat. Similar assessments conducted on other continents preceded and followed enactment of the Ramsar Convention on Wetlands of International Importance in 1971. In the United States, state wetland regulatory laws, starting in 1963, stimulated development of the first formal wetland assessment methodologies. These laws, now in effect in about 16 states, require individuals, private firms or public agencies to obtain permits before starting projects that would alter wetlands. Formal methodologies for assessment of wetland functions and values

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originated with a series of relatively simple models (Larson 1976) to assess wetlands in the state of Massachusetts, the first governmental unit to establish a wetland regulatory programme. These models were designed to assist community conservation commissions to make decisions about issuing permits to private individuals and firms, developers, and governmental agencies that wished to alter wetlands. In Massachusetts, a community conservation commission is a local governmental body consisting of citizens appointed to catalogue, plan and make recommendations for the protection, enhancement and use of natural resources within their city and town. The commissions have been given the legal responsibility of issuing wetland permits when a landowner wants to fill, drain or otherwise alter a wetland. Their decisions on whether to allow the proposed alteration, to allow alterations under specific conditions, or to deny the request may be appealed to a state agency. In other states, local commissions may serve in an advisory role to a state agency that issues the permits. The existence of these commissions is an arrangement that ensures that local people have an important role in management of the wetland resource. Assessment of wetland functions assists these local authorities to make better informed decisions about wetlands in the local context by informing them of the specific functions present in their wetlands. In 1972, the United States extended nationwide protection to wetlands and many of their functions through enactment of the Clean Water Act (for a detailed discussion of the practical significance of Section 404 of this Act, see Smith, Chapter 24). The US Army Corps of Engineers became the permitting authority when the Congress expanded the Corps’ jurisdiction over dredging and filling in of navigable waters to all of the ‘waters of the United States’. The US Environmental Protection Agency was authorised to set the standards for the national wetland regulatory programme. Other federal natural resource agencies were authorised to review permit applications and make specific recommendations to the Corps of Engineers.

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Methodologies for Wetland Assessment National regulation created an environment that led several federal agencies to sponsor development of wetland assessment methodologies that were expected to be applied across the country (US Fish and Wildlife Service 1980; Adamus 1983; Adamus and Stockwell 1983; Adamus et al. 1987). During the period 1976 to 1993, private consulting firms, and federal, state and provincial agencies in the US and Canada developed over 20 new assessment methodologies or variations on earlier methods (Larson and Mazzarese 1994). More recently Bartoldus (1999), for North America, and Finlayson et al. (2002), globally, have summarised the status of wetland assessment.

TRE ND S IN T HE E V O L U T ION OF WE T L AN D AS S E S S M E N T Since creation of the original assessment methodologies, the role of wetlands in providing the ecosystem goods and services on which human welfare depends has gained much wider public recognition. The total number of goods and services varies depending on whether a particular list combines certain functions or treats them separately. Assessment techniques have expanded in their scope in order to keep pace with the increased number of functions that have gained recognition. Natural resource managers are currently seeking to develop methods that can deal with the consequences of change that may be brought about, not only by direct human impacts on wetlands, but by changes in sea level, storm patterns and other impacts that derive from global climate change. The number of assessment methods has also been influenced by the fact that assessment techniques were initially constructed for (i) application to specific and limited areas, (ii) conformity with the specific provisions of different wetland permitting laws, and (iii) use with different legal definitions of wetlands. The evolution of wetland assessment methodologies has followed a pattern similar to that experienced in the evolution of wetland definitions.

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Early wetland definitions were developed to assist in wetland inventories for migratory bird habitat. The definitions tended to be based on biological and physical characteristics that were important to breeding, migratory and wintering activities of water birds. These characteristics varied among ecosystems and were not selected with consideration of other wetland functions. Definitions were in some cases influenced by the specific characteristics and availability of information such as aerial photographs and soil maps. In some cases, the decision whether to define wetlands on the basis of one or more characteristics (soil, hydrology, vegetation, wildlife use, etc.) was made by agencies or legislative bodies responding to very different local political influences. As in the case of wetland definitions, early assessment methods were keyed to local or regional characteristics that were later shown to be unsuited for use in other regions. New assessment methods were developed to meet the requirements of different regions. This resulted in a proliferation of inconsistent methods, which hampered efforts to conduct meaningful assessment across large ecosystems and catchments shared by several political units (see below). Consequences of use of inconsistent wetland assessment methods – a case history Each of the six small states in the New England region of the north-eastern United States has its own wetland regulatory programme. Different provisions in the wetland statutes in each state led to different wetland assessment methodologies despite the fact that these states share major catchments that cross their political boundaries. In 1989 the US Fish and Wildlife Service attempted to make an inventory of wetlands of prime importance in the Connecticut River catchment. Each of the four states that occupy parts of this large catchment was asked to list the most important wetlands in their portion of the catchment. But the results were not comparable across the catchment because one state had not adopted an assessment method and the

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three other states used different methods. Unless political subdivisions adopt compatible wetland assessment methods (Larson et al. 1998) it will be difficult to encourage consistent wetland management on a catchment or ecosystem basis. In recent years there have been increased efforts to develop wetland assessment methodologies based on fundamental biological, hydrologic and geographic characteristics that are widely applicable and not limited by differences in political jurisdiction. Nations seeking to benefit from the past successes and mistakes in wetland assessment would do well to focus on more recent methodologies.

T R AN S F E R AB IL IT Y OF M E T HODOL OGIE S Most formal wetland assessment methodologies have been developed to meet the requirements of governmental regulatory programmes in the United States. It is important to realise that these are structured to comply with provisions in the US federal and state constitutions that severely restrict the power of the federal, state and local governments to control private use of land. Unless a regulation has the explicit purpose of protecting health, welfare and safety or, on the national level, interstate commerce, it can be challenged as being an illegal taking of a person’s property without fair monetary compensation. Many wetland assessment methods in the US are closely tied to these legal restrictions. While some wetland assessment methods may address cultural, aesthetic and biodiversity functions of wetlands, these functions usually cannot be protected under US law unless the wetland is acquired as a gift or is purchased or leased from a willing landowner. It is important to recognise that the strong protection of individual property rights and the assignment of land use control to states and local communities in the United States differs from the situation in most other nations where land use regulation is a national responsibility. Nonetheless, it is widely recognised that the most successful natural resource conservation programmes are

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those that have effective local involvement and support. Recognition of local property rights and community values is key to developing local support for wetland management programmes and decisions. Wetland assessment methods that identify specific functions with specific wetlands can provide a better basis for understanding how wetlands relate to those rights and values within the local frame of reference.

T HE APPR OPR IAT E T IME FOR ASSESSMEN T : DUR IN G IN V EN T OR Y OR WHEN AN IMPACT IS PR OPOSED? Assessments can be made either as part of the initial inventory of the wetland resource or at the point when decisions have to be made about an impending impact. The usual practice in natural resource management is to make an initial assessment in conjunction with the initial resource inventory that is locating and delineating forest stands, soil types and wildlife habitat. Inventory assessments consist of collecting basic characteristics of different resource units in order to assign them to broad functional categories based on fundamental differentiating characteristics. These categories help establish the basic framework of a future management plan. Wetland assessment in the United States, however, is characterised by a different pattern, driven by the nature and timing of the regulatory approach. Assessment has tended to be applied primarily when the regulatory agency responds to a proposed impact. The absence of an inventory assessment that organises the resource into broad functional groups in advance of impacts tends to put a lot of expectations on impact assessment in an adversarial environment. This has prompted serious debate about the efficacy of the different kinds of wetland assessment methodologies (Kusler and Niering 1998). The US experience provides strong evidence that assessment in response to the appearance of an impact is a practice that should be avoided (Cwikiel 1988; Larson 1998). An understanding of how the regulatory approach has created this problem may be instructive.

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Methodologies for Wetland Assessment Public concern in the US about the effects that filling and dredging of coastal wetlands have on the decline of commercially valuable fish and shellfish populations led to enactment of the first wetland protection legislation. This occurred in 1963, when the legislature in the state of Massachusetts restricted development on marine salt marshes. This requirement led to an almost complete halt to salt marsh destruction in that state. Soon after enactment of the marine permitting process, concern about the effects on flood damage and water supplies of filling and draining freshwater wetlands led the Massachusetts legislature to extend the requirement for obtaining permits to inland wetlands. The permitting requirement extends protection to wetlands by requiring an individual, business or public agency to obtain a permit before being allowed to alter a wetland. Prior to the adoption of the first state regulatory programme in 1963, the only segment of the wetland resource that had been at all well catalogued and assessed was that of wetlands important to migratory waterfowl. The new authority to regulate wetlands was given to public agencies whose primary role was regulation and who had a minimal role in resource management. They viewed their charge to be protecting wetlands as individual cases when they were threatened with change. The responsibility to make inventories of wetlands was secondary. Thus wetland regulation in the United States has been characterised by piecemeal enforcement without the inventory and assessment that gives an overarching understanding of the resource. The major perceived needs of the agencies were a legal definition of a wetland, and procedures to determine the boundary of the wetland. The definition issue eventually grew to involve the highest political leaders of the nation (Environmental Defense Fund and World Wildlife Fund 1992). The size and configuration of a wetland are critical parameters because they set physical constraints on development projects that can be easily and quickly translated into monetary terms. Under this economic pressure, wetlands tend to be examined and assessed for their functions and values on a case-by-case basis with little

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understanding of the wider functional importance of the wetland in the catchment or larger ecosystem. Wetland maps were developed, not to locate and assess the resource as a whole, but to help managers and lawyers determine the end of the upland (terrestrial) and the beginning of the wetland. Assessment that is delayed until an impact is proposed results in debates primarily over boundary determinations and local economic issues, and ecosystem-scale issues tend to be ignored. Quantitative and economic measures need not be excluded from an initial resource inventory. The US Army Corps of Engineers happened to have a team in the Charles River catchment in Massachusetts when a major rain event struck. The engineers saw thousands of hectares of mainstream wetlands fill with floodwater. These broad level wetlands acted as detention basins that slowly drained for many days after the storm passed. As a result of the detention and slow release of the rainwater, the peak downstream flood crest was significantly lower and much less damaging than would have been the case had the wetlands been filled in or drained. The Corps of Engineers quantified this detention effect, naming it ‘natural valley storage’. They then compared the cost of purchasing the wetlands with the cost of installing traditional levees, dams and water control structures to achieve the same effect on the flood crest. It proved to be less costly to buy the wetlands than to pay for the structural alternatives (Doyle 1986). The Charles River is an example of quantitative assessment of a single significant function. In this instance, the flood control function was so critical to local interests in this catchment that it alone determined the fate of the wetlands. The assessment led to the protection of large wetland areas that are continuing to provide a wide array of other goods and services as well. But, for many if not most situations, funds will not be available to conduct a full quantitative assessment of all wetland functions. It will be unrealistic to expect that most assessments will be quantitative, and will identify the regulatory options in economic terms. It is more realistic to expect assessment

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to point to one or two functions that may warrant quantitative assessment. In most cases, however, assessment will provide relative measures of function. Assessment delayed to the point of impact presents the resource manager with several disadvantages. The manager is rarely able to control the timing of the assessment, and adequate time to do the assessment in appropriate detail and scope may not be granted because the developer requires a response before some predetermined time. The assessment also may be forced into the period of plant dormancy or ice and snow cover. In the case of functions that require a long time to get suitable hydrologic data, the resource manager may be hard pressed to obtain the necessary time if the developer claims that there was no advance notice that the site might require an extensive assessment. Waiting to conduct assessment on individual sites on an ‘as needed’, that is, non-strategic, basis results in wetland management decisions without consideration of the ecosystem or catchment scale. The assessment is done at a single site with little or no understanding of the relationship of the wetland to others in the catchment. Under this approach, the impact of the decisions on the wetland resource as a whole is poorly understood. Relying on site-by-site assessment does not provide advance information about wetland functions that might encourage a developer to consider the feasibility of alternative sites when he or she starts to plan a project.

T Y P E S O F AS S E S S M E N T METHODOLOGY In addition to the appropriate timing for conducting assessments, there are issues of time, effort and costs that are related to scale, and the desired output. Assessments conducted at the landscape level, based on small-scale inventory maps, will cost less and require minimal fieldwork, but produce general or relative results. Assessments of wetlands that have been mapped at a large scale can produce more specific information but can

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entail much more field effort, time and cost to produce specific, detailed and sometimes quantified results. Assessment methods tend to fall into a hierarchy: Tier I – Landscape assessment based on advanced inventory at small map scales: • assessment by category (predicts general probability of functions); • assessment of individual sites (probability of functions by rankings or scores). Tier II – Site-specific assessment at large map scales: • assessment based on detailed characteristics of individual sites (may be quantitative). Tier I – landscape, advanced or inventory-level assessment Landscape-scale assessment provides a preliminary and general indication of which wetlands are likely to have importance with regard to certain functions, values and risks. Methodologies for making this type of broad assessment may simply predict the presence of functions or may provide an estimate of the relative importance of the wetland, or category of wetlands, with respect to a given function. Assessing functions by categories of wetlands at the landscape scale The most simple and useful assessment is one that uses ‘rapid predictors’ (Adamus et al. 1987) based on wetland inventory location and size data, and brief field visits to place a wetland in functional probability categories. This can be a very important first step for developing nations. Some examples of assessment by functional categories at the landscape scale follow below (Larson et al. 1989; Larson 1995). Water quality maintenance function Brinson (1988) has suggested that the geomorphic setting of basin, riverine and fringe wetlands is

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relevant to the surface water quality maintenance function. Basin or depressional wetlands with small catchment areas and little water running through them have little opportunity to influence streamwater quality. Riverine wetlands along first and second order streams, high on the catchment, have both the capacity and opportunity to assimilate nutrients and affect water quality. The significance of these wetlands with respect to this function may vary depending on the level of nutrient loading (i.e. forest versus agriculture). Wetlands fringing third order streams and higher are small in comparison with the volume of water flowing through them and they have little impact on water quality. By examining the location and size of wetlands on an inventory map, wetlands can be sorted into these categories at little cost. This gives a first indication of which wetlands will merit investment of time and money in detailed assessment of the water pollution control function when an impact arises.

whether a wetland has potential as a groundwater recharge site. Larson et al. (1989) adapted these indicators in a training manual developed in China for use in developing nations. A similar adaptation was developed for use in India (Larson 1995). Potential recharge can be estimated by examining topographic and surficial geologic information. Potential groundwater recharge can occur at wetlands that have stream inlets and no outlets, at sites located at or just below the crest of a major hill or mountain, in wetlands located within the permanent pool of a reservoir, and in wetlands located on well sorted glacial sediment, alluvial sediment and on some types of volcanic material such as pumice. In arid and semi-arid regions, wetlands situated at canyon mouths, where both the flow of water and the wetland are intermittent, are likely to be recharge sites. At the seasons of high groundwater levels, some of these sites are subject to reversal, and become discharge sites.

Flood control

Groundwater discharge

Ogawa and Male (1986, 1990) demonstrated on catchments in eastern Massachusetts that inventory data can be used to place wetlands in categories according to their potential to influence downstream flood crests. The effectiveness of wetlands in reducing downstream flooding increases with (i) the area of the wetland, (ii) its location downstream, and (iii) the distance from the area subject to flood damage. The largest mainstream wetlands on fourth order, or higher, streams in the mid to lower reaches of the catchment will have the highest impact on downstream flood crests. Smaller wetlands bordering on third order streams, or higher, often in the upper reaches of the catchment, will be less important for flood control. Wetlands isolated from stream flow will have the least influence on downstream flood crests.

Similarly, the above cited authors suggested indicators, useful for developing nations, to assess whether a wetland was a site where groundwater discharges to the surface. This is the most usual relationship of wetlands to groundwater. Rapid indicators of this function are: wetlands with outlets but no inlets; the presence of obvious springs; those located at or just above the base of major hills or mountains; or whose early morning water temperatures are cooler than the air in the hot season and warmer than the air in the cool season. In arid and semi-arid regions, wetlands at mouths of canyons, where both the water flow from the canyon and the wetland is perennial, may be discharge sites.

Groundwater recharge Adamus et al. (1987) developed a series of rapid indicators for making an initial assessment of

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Relationships between the source of water in a wetland and other wetland functions Larson et al. (1998) suggested that analysis and diagrams of the flow components (overland flow, groundwater flows, precipitation) and geomorphic position of freshwater wetlands could be

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Ov

erla

nd flow

Evapotranspiration

Precipitation

(a) Surface water depression

Water table usually below wetland level

erla

nd flow

Evapotranspiration

Ov

Precipitation

(b) Groundwater depression

Water table

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used to predict the probability of the wetland performing an array of other functions. These flow components were first illustrated by Novitzki (1982) and expanded by Novitzki for Larson et al. (1998). For example, a wetland in a depression with no groundwater flow, and only overland flow (no inlet or outlet stream), is unlikely to provide fish habitat but is likely to provide habitat for other wildlife (Figure 21.1). A wetland on a slope at the edge of a lake (Figure 21.2a) is likely to provide a large array of wetland habitat and water quality functions. A wetland on an upland slope (Figure 21.2b) is unlikely to provide flood

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Fig. 21.1 Fish are unlikely to be found in a depressional wetland without inlet or outlet stream.

control, water quality or fish habitat functions. A wetland in a floodplain (Figure 21.3a) is likely to provide floodwater storage and wildlife habitat functions, whilst a wetland in a floodplain and fed by groundwater up-welling though an opening in a peat deposit (Figure 21.3b), is not likely to provide floodwater storage but is likely to be responsible for increased vegetative and habitat diversity. While the interactions between surface water systems and potential wetland functions often can be estimated in general field surveys, the flows of surface and groundwater in wetlands situated in depressions can be complex (Figure 21.4).

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Precipitation

Evapotranspiration

(a) Surface water slope

Lake or river floodwater

Overland flow

Water table usually below wetland level

Evapotranspiration

Overla nd flow

Precipitation

(b) Groundwater slope

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Fig. 21.2 A wetland on an upland slope (b) is unlikely to provide the flood control, water quality or fish habitat functions of a slope wetland bordering a lake (a).

To understand the relationships between groundwater flows and functions within, or down slope from, a wetland in a depression may require costly installation and monitoring of groundwater wells over an annual hydrologic cycle. However, knowing that the relationships shown in Figure 21.4 may be present will help to interpret observations made within and down slope from these types of wetland.

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Groundwater inflow

Wildlife habitat Limited habitat information for certain forms of wildlife habitat can be predicted from landscape inventory data. For example, wetlands that are isolated from streams, ponds and lakes, and thus free from fish predation, can be important breeding sites for amphibians. While these wetlands may be too small to appear on small-scale

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Precipitation

Evapotranspiration

(a) Surface water flat

Lake or river floodwater

Evapotranspiration

Precipitation

(b) Groundwater flat

Groundwater inflow

inventory maps, they are an identifiable category by description. The first wetland wildlife habitat assessment methods (Golet in Larson 1976; Golet and Davis 1982) assigned wildlife values to categories of wetlands at the landscape level. At the time that these methods were being developed, the best information on wetland wildlife habitat came from decades of research and experience on the relationship of open water and communities of vegetation to habitat requirements of breeding migratory waterfowl (Cowardin et al. 1979). These early methods incorporated the assumption that the general position of the wetland in

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Fig. 21.3 Floodwater storage is more likely in a wetland on a floodplain (a) than in a wetland fed by groundwater upwelling (b).

the landscape, along with interspersion and juxtaposition of open water with major communities of vegetation, was an index to the numbers of animals and diversity of wildlife species that could be present. This assumption was a major compromise across habitat requirements of very different species and is no longer recommended by Golet or Larson (personal communication). The habitat requirements of nesting migratory waterfowl can still be assessed usefully at the landscape level using earlier traditional assessment methods. But the year-round habitat requirements of wetland dependent mammals, amphibians and reptiles requires assessment of specific characteristics of

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Overland flow

Evapotranspiration

Precipitaion

Evapotranspiration

(c) Through-flow

Precipitaion

(a) Inflow or outflow

Overland flow Recharge (springtime) Water table Water table Groundwater inflow (summer, fall, winter)

Evapotranspiration

Evapotranspiration

Precipitaion

Overland flow

Precipitaion

(d) Through-flow with occasional streamflow out

(b) Inflow

Overland flow

Recharge

Groundwater inflow

Streamflow out Water table Water table Groundwater inflow

Groundwater inflow

Recharge

Fig. 21.4 Possible flow components of wetlands fed by ground and surface water.

individual wetlands and the associated upland. As opposed to the earlier assessment methodologies based on broad habitat characteristics, these assessments are based on examining specific wetland sites for the presence of specific biological and physical features known to be associated with specific species (Whitlock et al. 1994a,b). Cultural aspects of wetlands Where information is already available from collateral sources on the cultural, historic, recreational, aesthetic, educational, geologic or

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scientific significance of wetlands, these may define categories of wetlands that merit special attention. Many wetland assessment methodologies make identification of wetlands of special interest the first step in the assessment process. Anthropologists have demonstrated that the presence of water bodies, including wetlands, is important in explaining the patterns of early human settlement. While they have not developed specific procedures for assessing wetlands, Coles (1984, 1988) has recommended a screening process so that archaeologists can establish priorities for assessing wetlands for historic and archaeological remains.

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Assessing wetland conditions: impact of urbanisation and sensitivity to specific impacts In areas where land use mapping has been done, the amount of impervious surface in a catchment can be used to assess the impacts of urbanisation on a wetland (Hicks in Larson et al. 1998). Larson et al. (1998) suggest that the geomorphic location of a wetland (depression, slope, flat), its area and configuration can be used to estimate its susceptibility to direct physical impacts like poorly planned timber harvesting, filling and draining. They suggest that the same parameters can be used to assess the degree to which a wetland is susceptible to impacts on its local catchment (direct overland flow into the wetland) and impacts that occur generally on the larger contributing catchment (stream flow into the wetland). For examples, small (12 ha) fed from surface water or groundwater discharge are less likely to be at risk to local direct impacts. Wetlands in surface water depressions and slope or depressional wetlands fed by groundwater are highly susceptible to impacts on their local catchments. The risk factor for small wetlands to landscape catchment impacts rises with increases in the percent of impermeable surfaces of the catchment. Assessment of wetlands on a catchment basis Abbruzzese and Leibowitz (1997) have developed a synoptic method, using landscape indicators, for assessment of wetlands on an entire catchment basis. McAllister et al. (2000) have refined this method by applying it to assign wetland restoration priorities to catchments in the prairie pothole region of the north-central United States using selected functional criteria. The synoptic approach deserves more attention.

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Assessing functions of individual wetlands at the landscape scale If field visits can be made to selected wetlands that have been identified on a landscape scale inventory, additional functions can be predicted or assessed. An example of this level of assessment is the Tier I ranking type of assessment proposed by Larson (1998) for the New England states in the US based on the work of Amman and Stone (1991), Roth et al. (1993) and Amman et al. (1986). These functions include ecological integrity, flood control potential, water quality, water source relationship, fish habitat, aesthetic quality, recreational and educational access and historical significance. The output of the process is a non-numerical ranking of the significance of a wetland with respect to specific functions. This methodology involves relatively few predictors and can be accomplished without special expertise. One version of this type of assessment produces numerical scores that indicate the relative importance of the wetland to one or more functions (Amman and Stone 1991). The scores relate to scales that have been developed from reference databases. The level of confidence in the numerical score output depends on the credibility of the reference database. A serious drawback of this variation occurs when users of the method sum and average scores for different functions to generate a mean score that is then used to compare one wetland with another. There is no basis for assuming equivalency across wetland functions. Authors of scoring methods persistently warn against the practice, but users continue to do it. Disputes that have followed this misuse have generated unwarranted criticism of assessment in general. Private wetland consultants working for developers in the north-eastern United States, have often converted the strictly qualitative (high, medium, low) functional rankings of wetlands produced by these methods to numerical scores (1, 2, 3). The authors of the methodology had published a specific warning against numerical conversions in their manual because it established

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Methodologies for Wetland Assessment an unjustified arithmetic relationship between levels on the functional scale never suggested by the authors. The consultants further compounded their error by averaging the numerical values across all functions. This practice implied an unjustified equivalence in the importance of functions. As a result, the regional wetland regulatory office of the US Army Corps of Engineers decided to refuse to accept any wetland evaluation method in applications for federal wetland permits. The Corps staff could have chosen to reject applications that contained this malpractice, and required that the consultants adhere to the original methodology. By excluding all assessment methodologies from the permit review process, they excluded the benefits of controlled assessment that is open to replication, in favour of less objective and uncontrolled anecdotal speculation about wetland functions. An important feature of ranking and scoring methods is that they require some first hand inspection of the wetlands and their landscape setting. This can uncover important information that cannot be seen on maps and aerial photographs alone. A ranking assessment is an improvement over category assessment, because it provides the resource manager with the ability to look for alternatives when faced with a project to alter wetlands. Managers working on landscapes where wetlands are abundant may object to ranking assessments. Under those landscape conditions there is not time or money available to apply assessment of every function to every wetland, large and small. But inability to apply assessment to every wetland is mainly a problem of perception. If there is a sound reason for assessing every site, a relatively simple methodology can be selected and applied by volunteers who have attended a basic training course. Usually it is most effective to conduct a wetland landscape category assessment and then use the results to assign priorities to wetlands for more detailed ranking assessments. Two examples are illustrative. In the case of the Charles River (above) the large size and hydrogeomorphic location of the wetlands proved to be rapid indicators of wetlands

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that are likely to be involved in a single function of overriding importance. At the other end of the size spectrum, wetlands that consist of small temporary ‘vernal’ pools have long been known to provide essential breeding habitat for certain amphibians. Experience in the Massachusetts vernal pool certification programme has shown that trained volunteer high school students can effectively identify and register these small wetlands in the state’s wetland protection programme. Burne (2001) has demonstrated that colour infrared aerial photography can be effectively used to rapidly detect vernal pools in Massachusetts. On-site assessment of individual wetlands Site specific, detailed assessments typically require considerable time, money and personnel with specialised training. The regulatory environment in the US has prompted the expenditure of much effort to develop rapid, inexpensive methods to obtain detailed wetland assessments. This is unlike the usual approach applied to the inventory and management of other natural resources. A comparison with soil inventories and maps is instructive. Small scale soil maps and assessments produced at a landscape scale are extremely useful for planning purposes, even acknowledging that there are significant inclusions of other soils within the mapped soil units. When a specific site is being considered for a project or for a specific crop, it is understood that a detailed on-site, large-scale professional assessment and map will be needed. This two-tiered approach is a logical framework for responsible management of the wetland resource. Wetland landscape inventory and assessment for planning purposes needs to be better distinguished from on-site impact assessment. The first is designed to predict the latter. Certain landscape characteristics of wetlands make them candidates for intensive assessment. Examples are: wetlands upstream from drinking water reservoirs or abstraction points; wetlands upstream from sites that experience frequent flood damage; sites known to be used by endangered species, or that have documented scientific research, cultural,

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or historic records; and those that recharge subsurface aquifers. The potential for a wetland to have one or more of these characteristics, requiring a detailed on-site assessment, can often be detected in advance. Examples of on-site individual methods appear in the Tier II level regional scale assessment proposed by Larson et al. (1998) and the Adamus et al. (1987) WET methodology developed for use at the US national scale. Wildlife habitat Direct observation of wildlife in wetlands is difficult. Whitlock et al. (1994a,b) have developed models for detecting the presence or absence of suitable habitat for specific wetland-dependent amphibians, reptiles and mammals. These methods are based on field observation of key physical, biological and chemical habitat features for which there is a basis in published research. The end result is not a prediction of whether the animals are present in the wetland, but a determination of whether the important habitat elements to support them are present. The purpose of these models is to provide to wetland managers, in a concise manual, the expertise of species specialists that would otherwise have to be hired in the form of consultants. Crowley et al. (1996) have developed a similar method for wetland-dependent birds. These methods are limited by the available literature on the habitat requirements of individual species, and the models are very different for each species, however they reflect the best available science. This modelling approach can be applied to all species in any region. The US Fish and Wildlife Service (1980) has developed a method that predicts how many individuals of a species may be supported on a site. But this kind of assessment depends on knowledge of specific habitat requirements, which is not available for most wetland dependent species. Flood control Tier I landscape assessment distinguishes between wetlands that are most and least likely

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to provide the flood control function. This function can be quantified for specific wetlands only through professional engineering studies (Ogawa and Male 1986; Simon et al. 1987; Kittleson 1988). The Charles River example is an illustration of a quantitative and economic engineering study. Water quality maintenance Determining the quantitative role that a wetland plays in maintaining water quality is complex. It involves considerations of wetlands as sinks, sources and transformers of nutrients. The Wetland Research Program (1993) of the US Army Corps of Engineers has developed a screeninglevel assessment for the water quality function of wetlands that may be applied by professionals to sites where major impacts are anticipated. It is not likely that many studies will be funded to quantify this role for any but the largest wetland complexes. Condition or health of individual wetlands A recent development in individual wetland assessment is a method that assesses the health of a wetland by examining its biological integrity (Danielson 2002). Bio-assessment involves grouping wetlands into classes using existing classification systems and then selecting sampling sites across a gradient of human disturbances for each class of wetlands. These gradients range from sites that are minimally disturbed to severely degraded wetlands. A variety of proxies for human disturbance are used, such as percentage of the catchment in impervious surfaces or agriculture. Following assignment of the wetlands to classes, samples are then taken of at least two biological assemblages across the gradients. Commonly, assemblages include algae, amphibians, birds, fish, macroinvertebrates and vascular plants. From this sampling an index of biological integrity (IBI) is developed (Danielson 2002). Biological integrity is the ability of a wetland to support a taxonomic assemblage of organisms similar to that found in appropriate reference wetlands that have had minimal human impacts.

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Methodologies for Wetland Assessment A detailed examination of the biological attributes of each assemblage is conducted in order to detect patterns of response to human disturbance. Between 8 and 12 metrics are selected that show empirical and predictable changes across a gradient of human disturbance. These are eventually expressed as numerical scores that indicate the overall condition of the wetland. Developers of the bio-assessment approach are seeking to have public agencies incorporate the approach into water quality improvement programmes. The use of biological indicators of aquatic health is not new but, until recently, wetlands have not been viewed as part of the aquatic system for evaluation purposes. Macroinvertebrates have long been used as biological indicators of the quality of stream, pond and lake waters. But Hicks (1997) has shown that protocols and macroinvertebrate species that have been effective for assessing the condition of open waters of streams and ponds are inappropriate for assessing the adjacent wetlands, and has developed a macroinvertebrate protocol for assessing freshwater wetlands. Carlisle (1998) and Carlisle et al. (1999) have developed a similar protocol for coastal salt marsh wetlands. Relationships of wetlands to water sources Tier I assessment can be used to indicate whether there is the potential for a link to groundwater aquifers. Quantification of that link requires professional hydrogeologic study usually over a year or longer. This level of effort is usually limited to wetlands that are associated with major public water supply wells. Historical and cultural significance Assessment of the historical importance of wetlands requires professional study by field archaeologists. These studies start with initial surveys to determine the potential of the site. Based on the results of the initial survey, more extensive studies may be recommended. Coles (1988) has recommended a system for professional archaeologists to use for ranking wetlands for historical

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significance. Shiyam and Smardon (1990) have developed a detailed wetland heritage assessment methodology to be used as part of the US Army Corps of Engineers WET assessment process (Adamus et al. 1987). The hydrogeomorphic approach A relatively new approach to assessment is the hydrogeomorphic (HGM) method. This method is more fully treated elsewhere in this section (Brinson, Chapter 22; Maltby et al., Chapter 23, describe a European version). Mention here will serve only to put HGM in perspective with earlier methods. Early wetland classification (taxonomy) systems were developed as part of programmes to identify wetlands that were important for migratory waterfowl. This was the case for the first national wetland classification system in the United States, for the early inventories of wetlands of international importance, associated with the Ramsar Convention of Wetlands of International Importance especially for Waterfowl. It remains true for most wetland classification systems used in the world today. The life forms and spatial pattern of the wetland major vegetation and open surface water are key elements in these classification systems. The first wetland assessment methods were very much influenced by these same characteristics. Hydrology is the main forcing function of wetland ecosystems. The late Ted Hollis of University College London frequently reminded his colleagues ‘Hydrology put the wet in wetlands’. Hydrology, closely linked to the physical setting of the wetland in the landscape, is a robust basis for wetland assessment. Recent assessment methods developed on a regional scale have been evolving in this direction. HGM is an attempt to develop a broadly applicable method by combining these characteristics with the use of reference wetlands for calibration. Because hydrology is fundamental to understanding the functions of wetlands, HGM will probably exert a strong influence on the future development and refinement of wetland

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assessment. However, like many natural resource procedures developed in the industrialised world, it depends on the availability of map information, site data, and trained personnel that are commonly not available in developing nations. For this reason, wetland resource managers in many regions of the world will use wetland and landscape characteristics that can be economically acquired in a short period to make inventories and conduct a first screening for potential wetland functions. For specific examples of the application of the Hydrogeomorphic Approach to various ecosystems, see the US Army Corps of Engineers (Waterways Experiment Station, Vicksburg, Mississippi, USA) guidebook series (from 1993 to 2006) on this topic, and also Brinson (Chapter 22).

A SSE SS M E N T AS A M AN AG E M EN T TO O L F O R AL L OCAT IO N O F S TAFF A N D F IS CAL R E S O U R CE S Little attention has been given to the opportunity that assessment presents to strategically organise and allocate personnel and financial resources. Every natural resource manager has fewer people and funds than are needed to accomplish all the tasks that need to be done. Priorities have to be set and contingency plans developed to address unpredictable problems. An important role for wetland assessment is to provide the manager with a basis for allocating people, money and attention to threatened wetlands on an established and well understood priority basis. Assessment can put the manager in a position to develop a strategic first response plan to deal with a range of unexpected threats. This is not unlike advance planning for responses to human and natural disasters. With a response plan in place, a manager can use an assessment to decide whether the wetland in question needs only a desk review or requires a full scale on-site response involving a full suite of professional experts. It can be used to look at developing trends in impacts and to predict the level and types of resources that will be needed in future years.

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Landscape assessment can perform an important role in providing advance notification to agencies, firms and individuals of what to expect when they select the locus for their project, whether it is a new road, dam or housing development. If the results of landscape assessment are available for public inspection, developers cannot easily claim ignorance, or seek special treatment, on the basis that they had no reasonable way to know in advance the kinds of issues and studies they would face if their project were to impact wetlands.

WET LAN D ASSESSMEN T IN DEV ELOPIN G N AT ION S Assessment methods created in North America, Europe and other industrialised nations may require resources not available in developing nations. But initial landscape assessment does not depend on using expensive detailed procedures. The first step is an inventory of wetlands that includes a category assessment. An initial decision has to be made about which wetlands to include in the inventory. It is usually impractical to include all wetlands, especially on wellwatered landscapes. Some kind of cut-off needs to be set that makes sense in the local context. It can be a size threshold or it could be certain categories of wetlands that are of high current interest. The following are examples of the kind of information that needs to be noted at the time the first wetland inventory is made or when an existing inventory can be annotated with wetland information: • The general geomorphic nature of the wetland site: depression, slope, or flat (for diagrams see Larson et al. (1998) and Jaschke (1998)); • The primary source or sources of water: precipitation, groundwater or surface water; • The position of the wetland in relation to stream order. Is it isolated from surface water bodies, and if not, on what stream order is it located?; • If available, the type of surficial geology that underlies the wetland;

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Methodologies for Wetland Assessment • Any recorded or local anecdotal information on whether the wetland dries out every year, every few years, or never completely dries out; • Basic descriptors or sketches of the major vegetation types, their interspersion and juxtaposition; • Any special known resource (including food, fuel, fibre or fodder), historic, recreational, aesthetic, educational, geologic or scientific attributes of the wetlands. The accuracy and level of detail collected for the items above can reflect only the information at the time of the inventory and sometimes only general estimates will be available. Verma (2001) provides an example of how a basic ecosystem model of water quality parameters, categories of uses and impacts, identified local stakeholders, and several valuation techniques, can help policy managers in India assess wetland management options. The constraints on assessment in developing countries, such as problems of time, funding or incomplete data, are addressed by Roggeri (Chapter 25) who illustrates the use of a ‘functions and values matrix’. This method assigns values to functions according to their economic, social or ecological worth to the population, and can be used for the screening of interventions, environmental impact assessment, resource assessment, identification of stakeholder groups, zoning of land uses, cost-benefits analyses and feasibility of projects and the setting of priorities. In a case study, Roggeri cites the Amu Darya Delta wetlands adjacent to the Aral Sea in Uzbekistan (Ainsle and Sparks 1999; King et al. 2000).

CO N CL U S IO N S Wetland assessment has evolved over time. It initially involved only identification of important migratory bird habit and has progressed to formal and sometimes complex methods to assess a wide range of wetland functions including habitat for many wildlife species and functions important to human health and welfare. Most formal assessment methods have their origin in North America. Some aspects of these reflect unique legal conditions unsuited for

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other nations unless modified. It is important to note that while the United States has led in developing assessment methods it has, unfortunately, applied assessment when reacting to proposed impacts to wetlands rather than as a part of a strategic and holistic inventory of the status of the resource. Assessment that places wetlands in functional categories, at the landscape level and in conjunction with wetland inventory, is relatively rapid, inexpensive and does not require highly trained staff. It gives the resource manager an initial overview, at a catchment or landscape scale, of which wetlands are likely to perform what functions. This same overview can help resource managers design a strategic wetland management programme and response plan in anticipation of future impacts. Developing nations will probably find landscape scale assessment (Amman and Stone 1991; Jaschke 1998; Larson et al. 1998) a useful first step that will help identify which sites may later warrant application of more complex methods.

R EFER EN CES Abbruzzese B. and Leibowitz S.G. 1997. A synoptic approach for assessing cumulative impacts to wetlands. Environmental Management 21(3), 457–475. Adamus P.R. 1983. A Method for Wetland Functional Assessment, Vol. II, FHWA Assessment Method. Report Number FHWA-IP-82-24, Federal Highway Administration, US Department of Transportation, Washington, DC, 134 pp. Adamus P.R. and Stockwell L.T. 1983. A Method for Wetland Functional Assessment: Volume I, Critical Review and Evaluation of Concepts. Report Number FHWA-IP-82-23, Federal Highway Administration, US Department of Transportation, Washington, DC, 176 pp. Adamus P.R., Stockwell L.T., Clairain E.J. Jr., Smith R.D. and Young R.E. 1987. Wetland Evaluation Technique (WET), Vol. II, US Army Corps of Engineers, Waterways Experiment Station, Vicksburg, MS, 206 pages plus appendices. Ainslie W.B. and Sparks E.J. 1999. The hydrogeomorphic approach implemented: a Section 404 case study. Wetlands Bulletin 16(2), 8–9.

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Amman A.P. and Stone A.L. 1991. Method for the Comparative Evaluation of Nontidal Wetlands in New Hampshire. Pub. NHDES-WRD-1991-3, NH Department of Environmental Services, Concord, NH. Amman A.P., Franzen R.W. and Johnson J.L. 1986. Method for the Evaluation of Inland Wetlands in Connecticut. Connecticut Department of Environmental Protection. Bulletin No. 9. Ayensu E., van Claasen D., Collins M., Dearing A., Fresco L., Gadgil M., Gitay H., Glaser G., Juma C., Krebs J., et al. 1999. International ecosystem assessment. Science 286, 685–686. Bartoldus C.C. 1999. A Comprehensive Review of Wetland Assessment Procedures: A Guide for Wetland Practitioners. Environmental Concern Inc., St Michaels, MD, 196 pp. Brinson M.M. 1988. Strategies for assessing the cumulative effects of wetland alteration on water quality and others. Environmental Management 12(5), 655–662. Burne M.R. 2001. Massachusetts Aerial Photo Survey of Potential Vernal Pools. Natural Heritage and Endangered Species Program. Division of Fisheries and Wildlife, Westborough, MA. Carlisle B. 1998. Wetland Ecological Integrity: An Assessment Approach. Massachusetts Coastal Zone Management. Executive Office of Environmental Affairs, Boston. Variously paged. Carlisle B.K., Hicks A.L., Smith J.P., Garcia S.R. and Largay B.G. 1999. Plants and aquatic invertebrates as indicators of wetland biological integrity in Waquoit Bay watershed, Cape Cod. Environment Cape Cod 2(2), 30–53. Coles J. 1984. The Archaeology of Wetlands. Edinburgh University Press, Edinburgh, Scotland. Coles J.M. 1988. A wetland perspective. In: Purdy B. (editor), Wet Site Archaeology. The Telford Press, Caldwell, NJ, pp. 1–14. Cowardin L.M., Carter V., Golet F.C. and LaRoe E.T. 1979. Classification of Wetlands and Deepwater Habitats of the United States. US Dept. Interior, Fisha and Wildlife Service, Washington, DC. Crowley S., Welsch C., Cavanaugh P. and Griffin C. 1996. Habitat Assessment Procedures for WetlandDependent Birds in New England. Draft Report to the Federal Highway Administration and US Army Corps of Engineers. Dept. Forestry and Wildlife Management, University of Massachusetts, Amherst, MA. Cwikiel W. 1988. Looking for a reasonable return from assessment methodologies. National Wetlands Newsletter 20(3), 5–6.

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Danielson T.J. 2002. Methods for Evaluating Wetland Condition: Introduction to Wetland Biological Assessment. Publication No. EPA-822-R-02-014. Office of Water, US Environmental Protection Agency, Washington, DC. Doyle A.F. 1986. The Charles River watershed: a dual approach to flood plain management. In: Kusler J.A. and Riexinger P. (editors), Proceedings of the National Wetland Assessment Symposium. Association of State Wetland Managers, Box 528, Chester Vt. 05143, pp. 38–45. Environmental Defense Fund and World Wildlife Fund. 1992. How Wet is a Wetland? – The Impacts of the Proposed Revisions to the Federal Wetlands Delineation Manual. Environmental Defense Fund and World Wildlife Fund, Washington, DC, 175 pp. Finlayson C.M., Begg G., Humphrey C.L. and Bayliss P. 2002. Developments in Wetland Inventory, Assessment and Monitoring. Wetlands International, Wageningen. Golet F.C. and Davis A.F. 1982. Inventory and habitat evaluation of the wetlands of Richmond, Rhode Island. Occasional Paper in Environmental Science No. 1., College of Resource Development, University of Rhode Island, Kingston, RI, 48 pp. Hicks A.L. 1997. New England Freshwater Wetlands Invertebrate Biomonitoring Protocol (NEFWIBP). Natural Resources Environment and Conservation Program, University of Massachusetts Extension, Amherst, MA, 42 pp. Hollis G.E., Holland M.M., Maltby E. and Larson J.S. 1988. Wise use of wetlands. Nature and Resources 24(1), 2–13. Jaschke J. 1998. Minnesota Routine Assessment Method for Evaluating Wetland Functions (MnRAM), ver. 2.0. Minnesota Board of Water and Soil Resources. St Paul, MN, 44 pp. King D.M., Wainger L.A., Bartoldus C.C. and Wakeley J.S. 2000. Expanding Wetland Assessment Procedures: Linking Indices of Wetland Function with Services and Values. ERDC/EL TR-00-17, US Army Engineer Research and Development Center, Vicksburg, MS, 53 pp. Kittleson J.M. 1988. Analysis of flood peak moderation by depressional wetlands. In: Hook D.D., McKee W.H. Jr., Smith H.K., Gregory J., Burrell V.G. Jr., DeVoe M.R., Sojka R.E., Gilbert S., Banks R., Stolzy L.H., et al. (editors), The Ecology and Management of Wetlands, Vol. I. Timber Press, Portland, OR, pp. 98–111.

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Methodologies for Wetland Assessment Kusler J. and Niering W. 1998. Wetland assessment: have we lost our way? National Wetlands Newsletter 20(2), 1, 9–14. Larson J.S. (editor). 1976. Models for Assessment of Freshwater Wetlands. Pub. 32, Water Resources Research Center, University of Massachusetts, Amherst, 92 pp. Larson J.S. 1995. Identifying the functions and values of freshwater wetlands. In: Gopal B. (editor), Handbook of Wetland Management. Worldwide Fund for Nature, New Delhi, India, pp. 75–90. Larson J.S. 1998. Wetland landscape assessment: a way out of the dilemma. National Wetlands Newsletter 20(3), 3–4. Larson J.S. and Mazzarese D.B. 1994. Rapid assessment of wetlands: history and application to management. In: Mitch W.J. (editor), Global Wetlands: Old World and New. Elsevier, Amsterdam, pp. 625–636. Larson J.S., Nevel B.E., Whitlock A.L., Stevens T.H., Hicks A.L. and Jarman N.M. 1998. A Two-Tiered Approach to Freshwater Wetland Assessment in New England. Pub. No. 98-1, The Environmental Institute, University of Massachusetts, Amherst, 88 pp. Larson J.S., Adamus P.R. and Clairain E.J. Jr. 1989. Functional Assessment of Freshwater Wetlands: A Manual and Training Outline. Pub. No. 89–6. The Environmental Institute, University of Massachusetts, Amherst, 62 pp. McAllister L.S., Peniston B.E., Leibowitz S.G., Abbruzzese B. and Hyman J.B. 2000. A synoptic assessment for prioritizing wetland restoration efforts to optimize flood attenuation. Wetlands 20(1), 70–83. Novitzki R.P. 1982. Hydrology of Wisconsin Wetlands. Information Circular 40. University of Wisconsin Extension, Geological and Natural History Survey, Madison, WI. Ogawa H. and Male J.W. 1986. Simulating the flood mitigation role of wetlands. Journal of Water Resources Planning and Management 112(1), 114–128. Ogawa H. and Male J.W. 1990. Evaluation framework for wetland regulation. Journal of Environmental Management 30, 95–109. Ramsar Bureau. 2005. An Integrated Framework for Wetland Inventory, Assessment, and Monitoring (IF-WIAM). Resolution IX.1 Annex E on wetland inventory, assessment, and monitoring. 9th Meeting of the Conference of the Contracting Parties to the

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Convention on Wetlands (RAMSAR, Iran, 1971). Ramsar Secretariat, Rue Mauverney 28, CH-1196 Gland, Switzerland. Ramsar Convention. 1990. Guidelines for implementation of the wise use concept of the Convention. Annex to Recommendation C.4.10 (Rev.). In: Proceedings of the Fourth Meeting of the Conference of the Contracting Parties, Vol. I. Montreaux, Switzerland. Ramsar Bureau, Gland, pp. 179–182. Roth E., Olsen R., Snow P. and Sumner R. 1993. Oregon Freshwater Assessment Methodology. Oregon Division of State Lands, Wetlands Program, Salem, OR. Simon B.D., Stoerzer L.J. and Watson R.W. 1987. Evaluating wetlands for flood storage. In: Kusler J. and Brooks G. (editors), Wetland Hydrology. Proceedings of the National Wetland Symposium, Association of State Wetland Managers, Inc. Berne, NY, pp. 104–109. Shiyam C. A. and Smardon R.C. 1990. Wetland Heritage Assessment: Methodology and Literature Review as Part of Wetland Evaluation Technique (WET). Report 90-4, Institute of Environmental policy and Planning, SUNY College of Environmental Science and Forestry, Syracuse, NY. US Fish and Wildlife Service. 1980. Habitat Evaluation Procedures (HEP) Manual (102ESM). US Fish and Wildlife Service, Washington, DC. Verma M. 2001. Economic Valuation of Bhoj Wetland for Sustainable Use. Indian Institute of Forest Management, Bhopal, India. Wetland Research Program. 1993. Screening-Level Assessment of Water Quality Improvement from Wetlands. WRP Technical Note WQ-EV-2.1. US Army Corps of Engineers Waterways Experiment Station, Vicksburg, MS. Whitlock A.L., Jarman N.M., Medina J.A. and Larson J. S. 1994a. WEThings: Wetland Habitat Indicators for Nongame Species, Wetland Dependent Amphibians, Reptiles, and Mammals of New England, Vol. I, Pub. 94-1. The Environmental Institute, University of Massachusetts, Amherst, 45 pages plus computer disc. Whitlock A.L., Jarman N.M. and Larson J.S. 1994b. WEThings: Wetland Habitat Indicators for Nongame Species, Wetland Dependent Amphibians, Reptiles, and Mammals of New England, Vol. II, Pub 94-2. The Environmental Institute, University of Massachusetts, Amherst, 627 pp.

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22 The United States HGM (Hydrogeomorphic) Approach MARK M. BRI NSO N Biology Department, East Carolina University, Greenville, USA

IN T R O D U CT ION Functional assessment methods have been used in the United States for more than two decades to estimate the capacity of wetlands to perform functions. The most widely used method, the Wetland Evaluation Technique (WET), was spawned by Paul Adamus in the early 1980s (Adamus 1983). The technique was later modified with the support of the Waterways Experiment Station (Adamus et al. 1987), and has served as the basis for training and use by the US Corps of Engineers and other regulatory programmes. Beginning in the early 1990s, the HGM approach was developed to alleviate some of the shortcomings and dissatisfactions with WET (Smith et al. 1995). The approach, as it was initially developed, was designed to estimate positive or negative change in wetland condition using a quantitative comparison of altered or impacted wetlands with those that had not been altered. Effective application of the method would allow estimations of loss of condition in impacted wetlands and estimations of gain in condition in restored or created wetlands using ecosystem functions as the basis for evaluation. This would provide pertinent information for the public interest review process required by the US Federal Clean Water Act, Section 404 (33 United States

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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Code 1344). At the time of writing, the use of the HGM approach is not explicitly required as part of the impact assessment, minimisation, or mitigation phases of the Section 404 process. Although it was announced (Federal Register 1997) that the goal was to ‘develop, during the next two years, sufficient assessment models to address 80% of the Section 404 permit work-load requiring functional assessments’, funding was not allocated at a level necessary for developing regional guidebooks that could achieve this level of use. Several supporting publications for developing HGM guidebooks have been published, and at the time of this writing, 13 regional guidebooks are complete (Appendix). In spite of the lack of regional guidebook development for full implementation, three contributions of the HGM approach: the HGM classification, the use of reference systems and the distinction between wetland functions and values, have been influential in the evaluation of wetland condition in the USA. First, the HGM classification made it possible to distinguish between functions characteristic of different wetland classes. In so doing, the natural variation could be partitioned so that variation due to human impacts (alterations) would be more easily separated and measured. Second, reference wetland systems consist of sites spanning a range of conditions for each type, from the best to the worst conditions. Without reference, restoration and creation of wetlands had no agreed standards that were specific to different wetland types. Third, the lack of distinction between functions (from a scientific perspective)

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The United States HGM Approach and values (the goods and services supported by wetland functions for the benefit of society) caused inconsistencies to develop in the outcome of earlier assessments. By overlooking this distinction, assessments could generate conflicting values (say, bird hunting and bird watching) from the same function (e.g. maintenance of characteristic avian habitat). Societal goals for the protection of waters of the United States (which includes wetlands) are clearly articulated in the legislation that is the foundation of the Clean Water Act (33 United States Code 1344) from which national wetland regulation emanates. The goal of wetland assessments using the HGM approach is to determine the condition of a site, and a change in its condition due to human-

induced alterations, using the best science available. This provides decision makers with information to determine whether policy goals are being met as a result of proposed activities to change land use in wetlands. Wetland functions are the normal activities that take place in wetland ecosystems, or simply the things that wetlands do (Smith et al. 1995). Most functions provide the basis for the goods and services that contribute to socio-economic value (Figure 22.1). In contrast to values typically attributed to wetlands, functions occur whether or not humans are present to benefit or be influenced by them. By distinguishing between functions and values, the HGM approach provides a clear distinction between scientific information and policy deliberations (Figure 22.1).

Constructive feedback Natural energies: sun, water, nutrients

Societal infrastructure

Wetland ecosystem

Potential goods and services

Processes

Values Functions Consumers

Destructive feedback Natural stressors

Other ecosystems and fossil fuels

Fig. 22.1 Relationship between ecosystem functions and goods and services valued by society. The left-hand box represents natural wetland functions that can occur independent of society. The right-hand box represents the socioeconomic system that perceives wetland functions as goods and services. Values of consumers can generate activities that are either constructive or destructive to wetlands. (From Brinson and Rheinhardt (1998). Reprinted with permission from Lewis Publishers, Boca Raton, Florida, USA. © 1998.)

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Regulatory mechanisms available in the USA for wetland management are described in some detail in Smith (Chapter 24). Wetland assessments are used to facilitate decisions on individual projects that alter or eliminate wetlands. HGM assessments can be used in three ways: alternatives analysis, templates for restoration and creation and as standards for determining mitigation success. Alternatives analysis involves choosing which one of several project plans will cause the least environmental damage. As templates for restoration of particular classes of wetland, reference wetlands provide a blueprint for how to re-establish hydrology, manipulate the soil and introduce species. As a basis for judging success, the HGM approach uses mature and relatively unaltered wetlands as the standard. The approach is not designed to compare conditions between different wetland classes. This chapter provides an overview of the HGM approach to assessing wetlands, with ‘reference’ as the fundamental property that distinguishes it from other assessments. In so doing, the concept of reference standards (characteristics of relatively unaltered wetlands belonging to a wetland type) is introduced to provide managers with a benchmark from which to judge alterations to wetlands, including losses to other land uses. Information on reference wetlands, collected at field sites and derived from pertinent literature, can provide a tool for programmes that deal with impacts and restoration, and the net effect of these two processes.

THE CE N T R AL R O L E O F CL AS S IF ICAT IO N The primary objective of the HGM approach is to determine how alterations to wetlands affect their condition and their ability to perform functions. The challenge in making this determination is that natural wetlands vary widely in which functions they perform and the level at which they perform them (Brinson 1993a,b). The objective of the HGM classification is to identify a subset of wetlands of a particular type that function

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similarly and exhibit a relatively narrow range of variability under natural conditions. In so doing, variation due to impacts can be more easily identified and ranked according to intensity and effects on ecosystem functioning (Figure 22.2). The HGM classification was derived from one originally proposed for mangroves by Lugo and Snedaker (1974), and was expanded to cover all wetlands. The classification recognises seven types or generic classes: riverine, depressional, slope, lacustrine fringe, estuarine fringe and flats (mineral soil flats and organic soil flats) (Figure 22.3). Each of these geomorphic settings has characteristic dominant water sources and hydrodynamics. Riverine Riverine wetlands occur in floodplains and riparian corridors in association with stream channels. Dominant water sources are overbank flow from the channel for high order streams and occasional overland flow from adjacent uplands (nonwetland and low-lying areas) for low order streams. Additional sources may be interflow and return flow from adjacent uplands, tributary inflow and precipitation. When overbank flow occurs, the slope and roughness of the floodplain influences rates of flow. At their uppermost reaches, riverine wetlands intergrade with other wetland classes or uplands where channel (bed and bank) disappear. The downstream extent of riverine wetlands normally intergrades with estuarine fringe wetlands. Riverine wetlands normally extend from the stream channel to the upland edge of the stream’s floodplain, and include the stream itself. Perennial flow is not required. Riverine wetlands may lose surface water to the channel via overland flow after flooding and during rainfall. They lose subsurface water by discharge to the channel, movement to deeper groundwater (‘losing streams’, where the water table is lower than the stream) and evapotranspiration. Organic-rich deposits may accumulate in off-channel depressions (oxbows) that have become isolated from riverine processes and subject to long-term saturation from groundwater sources.

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Natural channel; buffered

Variation due to impacts

↑

↑

Assessment of condition

Not altered

Severely altered

Natural

Rural; unpolluted tributaries

Restoration Channelised; buffered

Dredged; leveed Degradation

Channelised; no buffer

Urban; nutrient - loaded

Dredged; polluted

1st–3rd order

>4th order

Beaver-dammed

↑

↑

Natural variation among classes HGM classification

Fig. 22.2 Distinction between classification, which partitions natural variation among wetland classes and assessment, which evaluates the effect of impacts relative to unaltered conditions. The classification axis separates subclasses of riverine wetlands according to stream order and influence by beaver dams. The assessment axis ranks subclasses by their departure from relatively unaltered conditions. Restoration and degradation occur within a subclass.

Bottomland (floodplain) hardwood forests and many riparian ecosystems (NRC 2002) are common examples of riverine wetlands.

discharge, they may slowly contribute to groundwater. Peat deposits may develop in some depressional wetlands.

Depressional

Slope

Depressional wetlands occur in topographic depressions. Prairie potholes of the upper mid-western USA are a common example. Dominant water sources are precipitation, groundwater discharge, overland flow and interflow (precipitation-derived subsurface flow) from adjacent uplands. The direction of flow is normally from the surrounding uplands toward the centre of the depression. Elevation contours are closed, thus allowing the accumulation of surface water. Depressional wetlands may have any combination of inlet and outlet channels, or lack them completely. Dominant hydrodynamics are vertical fluctuations. They may lose water through intermittent or perennial drainage from an outlet, or by evapotranspiration. Additionally, if they are not receiving groundwater

Slope wetlands are normally found where groundwater discharges to the land surface. Fens are a common example. They normally occur on sloping land; elevation gradients may range from steep hillsides to slight slopes. Slope wetlands are usually incapable of significant depressional storage because they lack the necessary closed contours. Principal water sources are usually groundwater return flow and interflow from surrounding uplands as well as precipitation. Hydrodynamics are dominated by downslope unidirectional flow. Slope wetlands can occur in nearly flat landscapes if groundwater discharge is a dominant source to the wetland surface. Slope wetlands lose water primarily by subsurface flow, overland flow and evapotranspiration. Slope wetlands may develop

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ne eri Riv

l

na

sio

s pre

De

e

pe

Slo

e rin

tua

Es

trin

us

c La

t We

t

fla

Fig. 22.3 Classes of wetlands showing geomorphic setting and the dominant direction of water flows. Solid arrows are surface flows; dashed arrows are groundwater flows. Dotted surfaces depict wetland areas. (From Brinson and Malvárez (2002). Reprinted with permission from Cambridge University Press. © 2002.)

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The United States HGM Approach channels, but the channels serve only to convey water away from the slope wetland.

intervening areas of low-elevation wetlands. Salt marshes and mangroves are common examples of estuarine fringe wetlands.

Lacustrine fringe Lacustrine fringe wetlands are adjacent to lakes where the water surface of the lake maintains the water table in the wetland. In some cases, they consist of a floating mat attached to land. Additional sources of water are precipitation and groundwater discharge, the latter dominating where lacustrine fringe wetlands intergrade with uplands or slope wetlands. Surface water flow is bidirectional, controlled by seiches (standing wave oscillations across the span of a lake) in the adjoining lake. Lacustrine fringe wetlands are indistinguishable from depressional wetlands when the size of the lake becomes so small relative to the fringe wetlands that the lake is incapable of controlling water tables. Lacustrine wetlands lose water by surface and subsurface flow returning to the lake and by evapotranspiration. Organic matter may accumulate in areas sufficiently protected from shoreline wave erosion. Unimpounded marshes bordering the Great Lakes of Canada and the USA are examples of lacustrine fringe wetlands. Estuarine fringe Estuarine fringe wetlands occur along coasts and estuaries and are under the influence of sea level. They intergrade landward with riverine wetlands where tidal currents diminish and river flow becomes dominant. At this interface bidirectional flows from tides dominate over the unidirectional flows of riverine wetlands. Water sources may include groundwater discharge and precipitation. Because estuarine fringe wetlands frequently flood, and water table elevations are controlled mainly by sea surface elevation, they seldom dry for significant periods. Estuarine fringe wetlands lose water by overland flow to tidal creeks (ebb tides) and by evapotranspiration. Organic matter may accumulate in higher elevation zones where flooding by tides is less frequent and where they are isolated from shoreline wave erosion by

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Mineral soil flats Mineral soil flats are most common on interfluves, extensive relic lake bottoms, or large floodplain terraces where the main source of water is precipitation. Pine flatwoods with hydric soils are a common example. They receive virtually no groundwater discharge, which distinguishes them from most depressions and slopes. Dominant hydrodynamics are vertical fluctuations. They lose water by evapotranspiration, overland flow and seepage to underlying groundwater. They are distinguished from flat upland areas by their poor vertical drainage, commonly due to spodic soil horizons and hardpans (which retain water), and slow lateral drainage, usually due to low hydraulic gradients. Mineral soil flats that accumulate peat can eventually become ‘organic soil flats’ (See Cook et al., Chapter 18). They are absent in arid climates. Organic soil flats Organic soil flats, or extensive peatlands, differ from mineral soil flats, in part, because their elevation and topography are controlled by vertical accretion of organic matter (See Cook et al., Chapter 18). They occur commonly on flat interfluves, but may also be located where depressions have become filled with peat to form a relatively large flat surface. Water source is dominated by precipitation, while water is lost by overland flow and seepage to underlying groundwater. Raised bogs share many of these characteristics, but may be considered distinct because of their convex upward profile and distinct edaphic conditions for plants. Portions of the Everglades, pocosin peatlands and northern Minnesota peatlands are examples of organic soil flat wetlands in the USA. While this classification serves as a conceptual basis for linking wetland type with groups of functions, it is not detailed enough to deal

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Table 22.1 Hydrogeomorphic classes of wetlands showing associated dominant water source, hydrodynamics, and examples of sub-classes (Brinson et al. 1995). Examples of sub-classes Hydrogeomorphic class

Dominant water source

Riverine

Overbank flow from channel

Dominant hydrodynamics

Eastern USA

Western USA & Alaska

Bottomland hardwood forests Carolina bays and vernal pools Fens Sapelo Island, Georgia

Riparian forested wetlands California vernal pools

Slope Estuarine fringe

Unidirectional and horizontal Return flow from groundwater Vertical and interflow Return flow from groundwater Unidirectional, horizontal Overbank flow from estuary Bidirectional, horizontal

Lacustrine Fringe Mineral soil flats Organic soil flats

Overbank flow from lake Precipitation Precipitation

Great Lakes marshes Wet pine flatwoods Peat bogs; portions of Everglades

Depressional

Bidirectional, horizontal Vertical Vertical

with variation that occurs within a specific geographic region. Several examples of subclasses are provided in Table 22.1. For example, floodplains of first and fifth order streams within a geographic region would both fall within the same generic class: riverine. For the purposes of functional assessment, however, the wetlands of the two floodplains may have such vastly different soils, vegetation and hydroperiod that it becomes impractical to assess them using the same standards. Thus, greater resolution can be gained by subclassifying riverine wetlands according to known properties of natural variation to further reduce the range of variability and increase the ability to detect the effects of human activities. In addition, some wetlands, especially if large, occur as interconnected mosaics of several subclasses that function differently from one another. Prior to conducting an assessment, individual subclasses should be identified so the appropriate assessment method is applied. If this is not done, the wrong set of standards will be applied, and a relatively unimpacted wetland could receive low scores, not because it was in a degraded condition, but because its class was wrongly identified.

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Avalanche chutes San Francisco Bay marshes Flathead Lake, Montana Large playas Alpine peat bogs

The need for regional subclasses is illustrated with two regional classifications: wet pine flats of the south-eastern coastal plain and riverine wetlands in the Piedmont region of the Carolinas and Georgia. This is done to show that: • the generic classification scheme is mainly a guide for illustrating differences in functioning rather than a strict set of rules for classification; • not all wetlands fit neatly into one of the seven generic classes; and • further subdivisions into regionally specific subclasses are needed to achieve homogeneity of natural features within a wetland type so that variation due to alterations can be more easily detected. Riverine wetlands of the Piedmont of the Carolinas and Georgia illustrate the need to generate several subclasses. These wetlands are in a state of transition due to changes in land use during the past 200 years (Trimble 1970). For example, the floodplains of some of these rivers continue to receive overbank flow as a dominant source of water on roughly an annual return period. Other floodplains have been isolated due to channel incision and widening, and now receive water mainly from lateral sources

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Class: Mineral soil flats

Class: Riverine

Hardwood flats Overbank flow dominated

Groundwater dominated

Beaver dam dominated

Bunchgrasspine

Wet pine flats

Cypresspine

Canebrakepine

Fig. 22.4 Classification of riverine wetlands in the piedmont of Georgia and the Carolinas and mineral soils flats in the coastal plain of south-eastern USA. (After Rheinhardt et al. 2002.)

(i.e. groundwater discharge and overland flow from adjacent uplands). A third type can be attributed to resurgence of beaver populations in the past three decades. The resulting beaver dams have raised water tables on many streams, thus fundamentally altering water velocity and depth behind their dams. This subclassification is warranted based on differences in fluvial geomorphology, hydroperiod and hydrodynamics (Figure 22.4). The mineral soil flats class is an example where the subclasses were not fully recognised until data were collected and a broad geographic perspective was gained. The south-eastern coastal plain of the USA (from eastern Texas, further eastward along the Gulf of Mexico coast, and then from northern Florida northward to North Carolina), was considered a potential reference domain for a subclass of wetlands variously called wet savannas, wet pine flats or pine flatwoods. After sampling approximately 70 sites and analysing data on mineral soil flats, Rheinhardt et al. (2002) developed three subclasses (Figure 22.4). They have sufficiently different plant community composition and fire frequency that it was necessary to separate them to develop relatively consistent reference standards that would apply across this large region. The bunchgrass/pine savanna is by far the most prevalent, the cypress/pine savanna is the wetter phase that differs in both canopy and ground cover species, and the canebrake/pine savanna (a somewhat provisional subclass) differs

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by dominance of Arundinaria gigantea in the understory. The major alteration to these subclasses is fire suppression, which allows shrubs and hardwoods to replace bunchgrasses and other shade intolerant plant species (Rheinhardt et al. 2002). In each of these subclasses, the generic classification provided the starting point and conceptual basis (geomorphic setting, hydroperiod and hydrodynamics) for the regional subclasses. However, subclasses were necessary to provide the specificity needed for functional assessment to resolve the effects of impacts. As a general rule, a balance must be achieved between classifying in so much detail that the approach is too cumbersome for practical application, and lumping in such large aggregates that the natural variation is too high to permit the detection of impacts.

R EFER EN CE WET LAN D SY ST EMS The use of reference in the HGM approach is central to developing an assessment method that uses existing, relatively unaltered ecosystems as the benchmark for comparison. In most cases, reference standards (defined below) represent conditions of the least altered sites of a geomorphic subclass. The purpose of these standards is to provide stability to the assessment by ‘referencing’ it to real wetland sites (Brinson and Rheinhardt 1996). Unaltered conditions have

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been used for many years in ecological studies to better understand the self-sustaining properties of ecosystems. Reference systems can be best explained by defining and discussing each of the component terms. As explained below, a reference system does not consist of a single site, or even a collection of unaltered sites. Instead, reference systems consist of the totality of information that is brought to bear on functional assessments, as embodied in the following terms. Reference domain All wetlands within a defined geographic region that belong to a single hydrogeomorphic subclass. Reference domain identifies boundaries where the subclass changes geographically, owing to a different climatic or physiographic region or a different biogeographic province. Transitions between reference domains may represent a continuum that is difficult to identify precisely. In practice, reference standards (defined below) of

(a)

a particular reference domain will be dissimilar to those in another reference domain, even if the wetlands are of the same class. Reference wetlands Wetland sites within the reference domain that encompass the known variation of the subclass. Reference wetlands are used to establish the range of conditions within the subclass. They may include former wetland sites for which restoration is possible. Historic records or published data may also contribute to the body of knowledge about the condition of sites that have been minimally or not altered. Typically, reference wetland sites are identified through extensive reconnaissance in the field with the aid of local experts, maps, aerial photographs, historic records and informed judgement. Figure 22.5 shows extreme conditions for a subclass of headwater stream, one in relatively unaltered condition (a) and the other severely altered by channelisation and replacement of forest by annual row crop (b).

(b)

Fig. 22.5 Two extremes in riparian and stream condition: (a) unchannelised stream with mature forested riparian zone in forested catchment (relatively unaltered) and (b) channelised stream with a riparian zone mostly in annual and perennial herbs in a row crop agricultural catchment (severely altered). The intermittent channel, now dry, passes vertically through the center of the photograph on the left. The photograph on the right is in the same subclass. (From Brinson et al. 2006.)

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Reference standard sites The subset of reference wetlands from which reference standards (below) are developed. Among all reference wetlands, reference standard sites are judged by an interdisciplinary team of local experts to be the least altered condition of the subclass, and hence display the most characteristic level of functioning. Characteristics of reference standard sites (e.g. reference standards) stabilise the assessment procedure by relating it to climatic and physiographic conditions found within the reference domain. The choice of reference standard sites is one of the most critical and controversial components of the HGM approach. The choices will influence the outcome of all subsequent assessments. If the subclass is too broadly defined (i.e. natural variation is large), or if degraded wetlands are included in the population of reference standard sites, then the resulting assessment procedure will lack the resolution necessary for detecting significant changes due to human activities (Figure 22.2). If the subclass is too narrowly defined there may be too few reference sites available, or the proliferation of subclasses could make an assessment programme unwieldy. Descriptors applied to these sites vary from ‘relatively unaltered’, which implies that they have been subjected to only natural disturbances with minimal anthropogenic influence,

to ‘pristine’, a condition that is unlikely to exist except in the most remote regions of the earth. Consequently, decisions about which sites are chosen as reference standard sites are somewhat subjective. That is why consensus of an interdisciplinary team of scientists is needed for choosing reference standard sites based on scientific literature, historic accounts, data on the sites and other sources of information. Reference standards Conditions exhibited by a group of reference wetlands that correspond to characteristic levels of functioning sustainable across the suite of functions of the subclass. By definition, reference standards, when combined as variables into the equations as described below, receive an index score of 1.0. In most cases, reference standards are equivalent to the measurements of variables (indexed to 1.0) from reference standard sites. The use of reference standards to scale indices of functioning is illustrated in Table 22.2. Some reference standards are present or absent (stream channelised; ditch across floodplain) while others are merely proportions based on the minimum of the range from reference standard sites (stand basal area, litter cover, volume of large downed wood). It is common for wetlands to undergo succession due to disturbances from fire, hurricanes and

Table 22.2 Field indicators, conditions in reference standard sites, and calculation of variable index for first and second order stream swamp forests. (Adapted from Rheinhardt et al. 1999.) Conditions for reference standard sites

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Field indicators

Range

Mean

Calculation of variable index (between 0.0 and 1.0)

Stream channelised Ditch across floodplain Stand basal area Litter cover Large downed wood (LDW)

No No 25–43 m2 ha−1 57–98% 2.1–18.1 m3 ha−1

No No 33 m2 ha−1 83% 9.5 m3 ha−1

Yes = 0.0; No = 1.0 Yes = 0.0; No = 1.0 BA/25 % litter cover/57 LDW/2.1

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interannual variations in hydrology. This dynamic nature of some wetlands can be captured by developing several sets of reference standards, each calibrated to the appropriate successional stage being assessed. This is to avoid scoring young successional stages low, when, in fact, they may be necessary component of a natural fire cycle, they are a consequence of channel meandering in a riverine wetland, or they represent a seral stage within a drought cycle. Because one of the goals of functional assessment is to separate natural variation, including natural cycles of disturbances, from alterations induced by human activity, the approach should make allowances for disturbances that sustain wetland ecosystems. In black spruce wetlands of interior boreal forests in Alaska, for example, unaltered conditions were identified for stand age categories of 0 to 5 years, 6 to 30 years, and greater than 30 years since burning (State of Alaska Department of Environmental Conservation 1999). Reference standard conditions were developed for each of these three successional stages. Site potential Level of functioning achieved under the least altered conditions given local constraints of disturbance history, land use or other factors. Site potential may be equal to or less than levels of functioning established by reference standards. It is especially important to recognise site potential in restoration projects so that reasonable standards of success can be established. Many restoration projects are carried out in urbanising landscapes where hydrology is undergoing change and habitat fragmentation is accelerating. Under such conditions, it is unreasonable to expect a restoration site to achieve reference standard conditions typical of relatively unaltered sites (Ehrenfeld 2002). Project target The final condition identified for a restoration or creation project. Conditions specified are used to judge whether a project reaches the target and is developing toward the site potential.

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While restoration projects should eventually achieve conditions equivalent to site potential, the bureaucratic time frame for judging success in the USA is normally no more than 5 years. Hydrologic conditions of a site may be well established at the beginning of a project, but vegetation and soil are slower to develop and difficult to evaluate, particularly for forested ecosystems that require many years to mature. Reliable professional judgement may be required in such cases to predict the composition and structure of a mature forest in comparison with one that was planted only 5 years ago. To this end, it would be appropriate to use successional stages that have a high probability of success as benchmarks for comparison. Project standards Specifications or criteria used to guide the restoration or creation site for purposes of compensatory mitigation. Project standards should specify reasonable amounts of intervention if the project target is not being achieved. Detailed descriptions of hydrologic modifications, planting schemes, weed control, soil preparation and so on, not only guide the party conducting the restoration, but provide a basis for project evaluation by regulatory agencies. Moreover, contingency plans should be outlined prior to project approval, to establish modes of response should the project fall short of the project target. Technical literature Pertinent information on the structure and functioning of the wetland type. Examples consist of ‘Community Profiles’ published by the US Fish and Wildlife Service (Wharton et al. 1982; Odum et al. 1984; Golet et al. 1993). Other worthwhile sources may be papers in scientific journals, edited volumes dedicated to specific wetland types and the ‘grey literature’ that may include research reports, government documents, environmental impact analyses and so on.

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The United States HGM Approach

Historical records Photographs, historic maps, land surveys and other documents. These materials may be useful in determining the unaltered condition, and may lead to useful information on the history of wetland impacts. There are alternative approaches to using relatively unaltered ecosystems as the basis for assessment. Such ‘non-reference-based’ approaches may enhance selected functions or may lead to restoration of wetland types that do not occur in nature. Such approaches fall outside typical HGM assessments where it is assumed that, by default, the goal is to manage the wetland for conditions equivalent to the relatively unaltered wetlands commonly found in a landscape. There are policy reasons for such alternative approaches. For example, sites can be enhanced to support waterfowl populations to meet societal demands for hunting, modified to optimise habitat for endangered species, configured to remove nitrate from agricultural runoff and planted to increase dominance of tree species that produce heavy mast. If such enhanced functions are to be maintained in the long run, intervention may be required, in contrast to a ‘hands off’ policy that prevails in the regulatory environment of the USA. One of the criticisms of the reference approach is that some landscapes have been so thoroughly altered that relatively natural ecosystems do not exist. For example, Ehrenfeld (2000) points out that many wetlands in suburban areas of New Jersey have modified hydrologies, excessive nutrient loading, invasive species and other types of alterations. Under such conditions, site potential is below that of reference standards. At least two alternative approaches can be taken. One is to use historical data for estimating the conditions of relatively unaltered wetlands. While such conditions could not be achieved nowadays, historical reference standards provide a basis for estimating how much the highest achievable condition (site potential) has shifted from a previously existing reference standard condition. A second approach is to set policy goals that are consistent with ‘desirable’ wetland types. Particularly in

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urban landscapes, many constraints may cause site potential to fall short of reference standard conditions. Wetlands under such novel circumstances may require maintenance activities such as replanting, erosion control, culling of invasive species and so on. The process of building a reference wetland system for a subclass is implied in the list of reference terms described above. For example, ‘reference domain’ implies that sites chosen for the subclass be restricted to a region where biogeographic differences in species composition are minimal. Reference standard sites should encompass as much of the natural variation as possible. To better define the transition to another subclass, a few sites could be characterised that fall just outside the intended subclass. This provides yet another frame of reference by indicating what the subclass is not. In addition to the approximately 20 sites for reference standard sites, an equivalent number of sites should be chosen that represent alterations due to common or significant impacts. Some sites are so severely impacted, however, that it would be trivial to actually sample them. However, impacts that lower the condition of the wetland, but do not eliminate it, provide useful information on indicators and variables that are the most sensitive to alteration. Much of the cost of data collection is travel so, once on site there are economic advantages to collecting more than the minimum anticipated data needed for functional models (described below). Two principal uses for more complete data sets are: as an aid to design of restoration projects that require detail beyond that needed for conducting a functional assessment, and use in updating models to make them more responsive to existing and unanticipated impacts. As a practical matter, some thought should be given to the expertise and training of the group responsible for data collection, and for model refinement and scaling. An early suggestion was to have regulatory agencies collectively gather data on reference sites and analyze them (Smith et al. 1995). This team approach has not proven very successful in some instances because regulatory personnel seldom have time to devote to such

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activities and are seldom trained to collect and analyse data. Qualified consultants and academics who have more flexible work loads can devote time and expertise to developing reference systems.

FU N CT ION S AS CUR R E N CY FO R E S T IM AT IN G E COS Y S T EM CON DIT IO N Functions are widely recognised as a way to deal with a subset of ecosystem attributes and processes, and to avoid the need to confront the complexity of the ecosystem as a whole (Maltby et al. 1994, 1996). Ecosystem functions tend to fall into three generic categories: maintenance of hydrology, biogeochemistry and habitat. Functions have been widely used historically for the assessment of wetlands (Larson and Mazzarese 1994). Functions are not essential or, indeed, the only way to estimate the condition of wetlands relative to reference. Such estimates could be accomplished using commonly measured community and habitat characteristics such as tree basal area, species composition and connectivity with adjacent forests. However, these measurements, without ecological interpretation, do not translate directly into ecosystem functioning. The advantage of aggregating variables into functions is to present information in a format that is easily understood and readily accepted. A further justification of using functions is their utility in relating them to societal goods and services. Most wetland management programmes require some level of public support. A framework that shows how human welfare benefits from ecosystem functioning may facilitate such support (Figure 22.1). Selection of functions Prior to selecting functions for a particular wetland subclass, the relevant research literature should be reviewed to establish how the wetland subclass works. Unfortunately, extensive research has seldom been conducted on particular regional subclasses. Consequently, professionals

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familiar with the relevant scientific literature, or those who have conducted research on similar wetlands, should become involved in choosing functions. Local experts with field experience should also be consulted. The choice of functions should be relatively broad, and relevant to the purpose of the assessment. A function like the cycling of nutrients is broad because it applies to all ecosystems and all essential elements. It is also relevant because it contributes to the maintenance of water quality, a condition that is consistent with the Clean Water Act of the USA. Other functions may be fundamental but not perceived as relevant. For example, the function of sequestering atmospheric carbon dioxide in peat-swamps and other wetlands with organic rich soils has global implications (Armentano and Menges 1986; Aselmann and Crutzen 1989). However, the influence of any one wetland site on atmospheric carbon dioxide would be minuscule and, as such, the function may be relevant only when placed in the context of cumulative impacts at large spatial scales (Wofsy 2001). It may be useful to incorporate functions that are specific to particular wetland subclasses, but not necessarily common or relevant to all of the subclasses in one of the seven classes. For example, headwater riverine floodplains of the mid and south Atlantic coastal plain remove nitrogen and phosphorus generated by land use activities, such as farming and urbanisation (Whigham et al. 1988). In such cases, nutrient removal may be a function of greater practical importance than nutrient cycling, and better tailored to the characteristics of the subclass. Table 22.3 provides lists of functions that were developed for low gradient riverine, prairie pothole and wet pine flat wetlands. Regardless of the functions chosen, all available information should be brought to bear in developing models that express the function, and ultimately condition. For wetland subclasses lacking research, one must rely on general information and the judgement by experts. In such cases, models of functions represent hypotheses with no explicit testing. This does not lessen the need, however, to make management decisions

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The United States HGM Approach Table 22.3 Comparisons of functions among three HGM subclasses. Class: riverine Subclass: low gradient riverinea

Class: depressional Subclass: rainwater basin in Nebraskab

Hydrology

Temporarily store subsurface water Maintain characteristic subsurface hydrology

Water storage

Maintain characteristic water level regime

Biogeochemistry

Cycle nutrients Remove and sequester elements and compounds Retain particulates Export organic carbon

Cycle nutrients Remove, convert, and sequester elements, compounds, and particulates

Maintain characteristic biogeochemical processes

Habitat

Maintain characteristic plant community Provide habitat for wildlife

Maintain habitat characteristics for plant community Provide wildlife habitat

Maintain site-quality for characteristic plant communities Maintain site-quality for characteristic animal communities

Category of function

Class: mineral soil flats Subclass: wet pine flatsc

aAinslie et al. (1999); bStutheit et al. (2004); cRheinhardt et al. (2002).

on wetland resources. Thus, the HGM approach is not a substitute for professional judgement, but rather a consistent framework within which to utilise and apply professional judgement. Development of models of functions, variables and indicators Once functions are identified, the variables that contribute to a function are determined. Variables are attributes of a wetland that contribute to a function. They become components of logic models or equations that depict a function, as described below. Continuing with the example of ‘nutrient cycling’, variables that contribute to the function include biomass components that accumulate nutrients through growth and that recycle nutrients through decomposition. Examples are living and dead biomass (e.g. density and size of trees, volume of large downed wood – e.g. fallen limbs and whole trees – leaf litter amounts, etc.). Field indicators that are proportional to variable conditions should be used to establish the range over which the variable fluctuates (Figure 22.6). Reference wetlands, described in the next section,

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are sampled to determine these ranges. A major assumption in the HGM approach is that unaltered conditions (i.e. largely unaffected by human activity) have the appropriate (and sustainable) level of functioning for the subclass. Variables are assigned a maximum score of 1.0 when they match reference standard conditions and 0.0 when the variable is absent. Between the 0.0 and 1.0 scores is a range of variable expression or condition. Functions may be represented by a single variable or a group of variables. Equation 22.1 shows how various biomass components can be combined into equations to depict a function. Consistent with the rapid assessment approach of HGM, variables (where Vi is the index for a variable) are chosen that are easily measured surrogates for biomass or other structural features of an ecosystem rather than direct measures of the function. For example, tree basal area (VTBA) represents the standing stock of tree biomass, shrub cover (VSHRUB) represents the cover of shrubs in the understory, and ground cover (VGROUND) represents the cover of herbaceous plants on the forest floor. Likewise, large downed wood (VLDW) is proportional to the

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Indicators (field/map measures)

Variables (scaled 0.0 –1.0)

Function (indexed to reference standard sites)

1 V1

2

3 V2

Index of function = [(V1 + V2)/2 + V3]/2

4 V3 5

Fig. 22.6 Relationship between indicators, model variables and indices of function. Indicators are typically map or field measurements (stand age, overbank flow evidence, species composition etc.), variables are ecosystem attributes that contribute to a function, and the index of the function represents the condition of the ecosystem as estimated by the function.

amount of woody detritus undergoing decomposition, number of stages of decay (VDECAY) represents the number of stages of decomposition of the large downed wood, and leaf litter (VLITTER) is proportional to the accumulation of litter on the forest floor. Nutrient Cycling Function

{

= (VTBA + VSHRUB + VGROUND ) / 3 + (VLDW + VDECAY + VLITTER ) / 3

(22.1)

}2

Equation 22.1 assumes that if major components of live and dead biomass are present, then nutrients will be cycled (e.g. through uptake in the production of biomass and release from decomposition of detritus). In short, the totality of the living and detrital biomass is assumed to correspond to the stock of elements that are currently participating in nutrient cycling through uptake, release, or storage (Brinson et al. 1995). If any of these variables

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depart from the levels characteristic of unaltered sites (either more or less), then the index of functioning should fall below the 1.0 level. This format is similar to that of the Habitat Evaluation Procedure (US Fish and Wildlife Service 1981). This same equation can be reconfigured to represent a nutrient removal function that captures the buffering effects of riparian zones for maintaining water quality by adding three variables: channelisation (VCHAN = presence or absence of a channelised stream); buffer condition (VBUFF = condition of vegetation buffers within 100 m of each side of the stream channel); and altered delivery of water to the wetland (VDITCH = presence or absence of a cross-floodplain ditch). Nutrient Retention Function  V + VBUFF + VDITCH  =  CHAN  3 

(22.2) 1/ 2

 × ( Nutrient Cycling Index )  

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The United States HGM Approach Instead of simply averaging the two groups of variables as for Equation 22.1, Equation 22.2 uses the geometric mean of the hydrologic and biomass components of the equation. Thus, if either of the components falls to zero, the function is considered to be totally eliminated. This approach has been applied to headwater riverine wetlands of the Atlantic coastal plain (Rheinhardt et al. 1999). In building models, variables may be weighted to reflect their relative importance to a particular function. Without specific data on nutrient uptake and release, however, assumptions should be stated on how weighting is justified. The simplest approach is to average variables and not weight them. Considerable research may be necessary to empirically test these relationships. Until such validation is possible, model construction must rely on related research conducted in other ecosystem types. As with other assessments, application of the HGM approach will measure condition before an impact, estimate condition after the impact, and use the difference between the two to estimate the loss. (Gains, rather than losses, are estimated for restoration projects over time.) Changes in functions are estimated through changes in relevant variables. As an example, VTBA (tree basal area) of unaltered sites may average 30 m2 per hectare (ha−1), which would be assigned a score of 1.0. An altered site may have a VTBA of 10 m2 ha−1, which, as a first approximation, might be indexed at 0.3, or approximately one-third of unaltered condition, assuming a linear relationship between the variable and its contribution to the function. Indexing is repeated for each variable and substituted in the equations for calculating the index of function. Care must be taken to avoid the assumption that ‘more is better’. For example, a litter layer that is four times the thickness of reference standard conditions is a signal that reference standard conditions are not met, perhaps as a result of low litter quality (for decomposers) or insufficient fire frequency. The exact variable index assigned in such cases may be somewhat arbitrary. However, few would argue that nutrient cycling in a stand with excessive leaf litter

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would be the same as an old-growth stand with a fraction of that amount. The equations are always configured mathematically to be indexed to the relatively unaltered sites. Discrete categories for variables can be used, with the simplest being 1.0 and 0.0 (present or absent). For example, if drainage transforms a wetland site to a non-wetland, one could argue that wetland-like hydrologic functions no longer exist, and the index of function for hydrology would be 0.0. More categories (e.g. 1.00, 0.75, 0.50, 0.25, 0.10, 0.00) can be employed if additional data become available to relate the structural variables measured to ecological condition. Alternatively, the linear example given above could be further calibrated by determining the response curve of the variable to an independent measure of ecosystem condition. Consistency among users is critical for effective application of the method (Whigham et al. 1999). In summary, the use of functions to estimate ecosystem condition includes a number of activities: • synthesising available information for the wetland subclass; • constructing models that depict functioning; • collecting relevant data from reference wetlands to scale the variables and ultimately the models; and • revising models based on the kinds of information that can be practically gathered during a rapid assessment procedure.

PR OCEDUR E FOR DEV ELOPIN G A R EGION AL GUIDEBOOK Table 22.4 identifies phases and specific steps for developing a regional guidebook. The first phase consists of identifying a wetland subclass of high priority for guidebook development. Priorities might be established because, for example, a majority of wetland impacts occur on the wetland subclass, the subclass is especially valuable and perhaps rare, or some other reason. At the same time, the reference domain can be provisionally identified. One approach is to start small

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Table 22.4 Steps in development of regional guidebooks. (Modified from Federal Register 1997.) Phase 1: Identification of Regional Wetland Assessment Needs A. Identify regional wetland subclasses B. Prioritise regional wetland subclasses C. Define provisional reference domains Phase 2: Draft Model Development A. Review existing models of wetland functions B. Initiate literature review C. Identify functions for each subclass based on literature reviews and expert opinion D. Identify variables and measures based on literature reviews and expert opinion E. Identify major impacts within reference domain and effects on functions F. Develop provisional functional indices G. Identify variables and indicators for sampling, and potential reference sites Phase 3: Sampling and Data Analysis A. Find reference wetland sites B. Collect data from sites C. Analyse data for determination of reference standards and scaling variables Phase 4: Model Calibration and Write Regional Guidebook A. Develop and calibrate models calibrated on data from reference sites B. Field test accuracy and sensitivity of functional indices C. Write guidebook Phase 5: Draft Regional Guidebook Review and Publication A. Obtain peer review of draft guidebook B. Conduct interagency and interdisciplinary workshop to field test and review guidebook C. Revise guidebook to reflect recommendations from peer review and workshop D. Obtain second peer review from scientific, consultant, and regulatory communities E. Publish guidebook Phase 6: Implement Draft Model Guidebook A. Identify users of HGM Functional Assessment B. Train users in HGM classification and evaluation C. Provide assistance to users Phase 7: Revise Regional Guidebook A. Determine need to subclassify further or expand reference domain B. Evaluate literature for improving models of functions C. Collect additional data, as needed D. Revise or supplement regional guidebook

to ensure that natural variation within a subclass is low, and not due to climatic factors. This leaves the option of expanding the geographic range in the future. This was the approach taken for low-gradient riverine wetlands in western Kentucky (Ainslie et al. 1999), precipitationdriven wetlands on discontinuous permafrost

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in interior Alaska (State of Alaska Department of Environmental Conservation 1999), and the depressions of the Rainwater Basin in Nebraska (Stutheit et al. 2004). In contrast, the wet pine flats effort began with a large geographic region (Rheinhardt et al. 2002) that now can be further subclassified if users determine that the

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The United States HGM Approach guidebook is not sufficiently specific for particular plant taxa as they vary regionally. Model development is an opportunity to identify and interpret pertinent literature. Regional experts who already know the literature and are familiar with reference sites would normally conduct this phase. It is also valuable to identify major impacts on the subclass in order that the development of functional indices can be structured to be sensitive to such alterations. Local knowledge of potential reference sites, especially relatively unaltered sites, is valuable because it saves time. Data collection and data analysis approaches can be found in previously published regional guidebooks (http://el.erdc.usace.army.mil/wetlands/guidebooks.html) as well as peer-refereed publications that have characterised subclasses in such a way that the natural variability is determined (Magee et al. 1999; Rheinhardt et al. 1999; Shaffer et al. 1999; Shaffer and Ernst 1999). Techniques for data collection are not unique to the HGM approach, but rather are common ecological sampling methods for forests and grasslands with emphasis on hydrologic indicators. Various types of multivariate analyses may be useful to characterise major factors responsible for variation within the subclass. The PC-Ord family of programs is commonly used for this purpose in community ecology (McCune and Mefford 1995). Ideally, separation of sites would be due to a single type of alteration as it affects a specific function. In reality, such separation is not always so ‘clean’, and calibration must be handled according to the best understanding of the relationship between impacts and effects on functions. At the core of a regional guidebook is a ‘profile’ of the subclass, which contains a review of the literature and sets the stage for a sciencebased assessment. This should precede sections on model description and the components of the guidebook that will be taken to the field for on-site assessments. The guidebook itself has components that are generic to all guidebooks for

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the purpose of standardising the methodology. However, in terms of how indicators, variables and models are developed, much flexibility is encouraged to make the best use of information available for the subclass. Peer review, including field-testing of guidebooks by user groups, can vary according to regional and national programme needs. If guidebook development is funded by a particular agency, that agency may want to exert quality control to ensure that the results from its application meet the needs of the agency’s wetland management programme. Once the peer review and field testing are complete, the final draft becomes available for training. Training courses should always be centred around a field approach because HGM assessments at the resolution described above require on-site data collection. The open architecture of the guidebooks encourages change and modification for improvement. As new scientific information becomes available, and as users of the guidebook reveal a need for improvement, revisions should be made in a timely manner.

APPLICAT ION OF T HE MET HOD The HGM approach can be divided into two phases: development and application. The foregoing discussion dealt primarily with the development phase that culminates in a guidebook for a subclass. While this work is formidable for each subclass, the intent is to minimise the time to conduct a rapid assessment and to maximise precision and accuracy. As currently designed, a HGM assessment can be conducted on a relatively routine project in one day or less. A routine project may consist of several hectares or less, an impact on a single hydrogeomorphic subclass, and comparison of two or three project alternatives to determine which is the least environmentally damaging. Projects with a restoration component would require more time and planning. Assessments can be done in less detail than described above, such as a cursory review of

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projects. An abbreviated assessment could take the form of a checklist of the probable effects of the project on the list of functions for the hydrogeomorphic subclass. At this level, the assessor could make provisional judgements on which functions would be affected and whether they would be partially or wholly eliminated. For some purposes, this may sufficient. At the other extreme, intensive data collection on hydrology or nutrient cycling may be required as part of a larger environmental assessment programme. The current design does not provide enough detail to quantitatively estimate individual processes; rather, it only estimates the deflection from reference standards based on condition. To illustrate how a typical assessment might be conducted, a hypothetical project is presented. Consistent with the process for reviewing potential impacts to wetlands, it is assumed that the project has high social significance and that compensatory mitigation will be used to offset environmental degradation. It is also assumed that the project has satisfied many other requirements that are not related specifically to wetlands, such as critical habitat for species of special concern, status under the Ramsar treaty, hazardous waste status and so forth (Smith et al. 1995). In this hypothetical example, the permit application is

being prepared with the aid of functional assessment using the HGM approach. For simplicity of explanation, the imaginary wetland subclass has only 5 functions (‘A’–’E’). All functions are indexed relative to 1.0, conditions derived from reference standard sites (Table 22.5). For illustrative purposes, all functions will be driven to zero for the project wetland (a highway project), so the loss in functioning is represented by the difference between pre- and post-project indices of function (e.g. −0.5, −0.9, etc., Table 22.5). Were it not for the social significance of the project, the permit might be denied. A site for restoration is identified, it is assessed as it exists before the restoration, and it is assessed again by projecting improved conditions anticipated after restoration. The indices for the restoration wetland are given in the right-hand portion of Table 22.5. These data depict a poorly functioning wetland before restoration that has potential for improvement. During the 5-year period when conditions of the restoration site improve through compensatory mitigation, the five functions of the restoration site are projected to increase in levels between 0.0 and +0.5. For function A, the gain in functioning from 0.2 to 0.7 (+0.5) is the same magnitude (−0.5) as lost by the project. Thus, a 1:1 mitigation ratio would

Table 22.5 Hypothetical example of wetland functions lost due to an unavoidable impact, and corresponding functions gained in a 5 year period by restoring a degraded wetland. Note that 5 years is insufficient time for functions to recover to a level of 1.0, the benchmark established from reference standard wetlands. (Modified from Brinson 1996.) Project wetland

Function assessed A B C D E

Index before project

Index after project

0.5 0.9 0.7 0.1 0.2

0.0 0.0 0.0 0.0 0.0

Difference pre–post project (losses) −0.5 −0.9 −0.7 −0.1 −0.2

Restoration wetland Index before restoration 0.2 0.3 0.1 0.2 0.2

Index after restoration

Difference pre–post restoration (gains)

0.7 0.6 0.1 0.5 0.7

+0.5 +0.3 0.0 +0.3 +0.5

Mitigation Ratio (area of restoration needed to replace a unit of loss) 1:1 3:1 NPa 1:3 2.5:1

aNot possible to restore this function under the conditions of the hypothetical restoration plan.

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The United States HGM Approach represent functional replacement, hectare for hectare. Correspondingly, function B requires a 3 : 1 mitigation ratio (0.9/0.3). Functional replacement is not possible for function C because there was no gain in function by the restoration wetland. For function D, the mitigation ratio is 1 : 3, meaning that only one-third of a hectare of restoration would replace the function lost in a hectare of project wetland. The last function, E, requires a 2.5 : 1 replacement. Table 22.5 represents the output of a hypothetical functional assessment. It does not provide guidance on how the results should be handled. As with any science-based assessment, policy is needed to handle the output. The example was construed to make several points. The first is that HGM functional assessment ends with the calculation of replacement ratios. Only three of many possible options are presented here on how to handle the output as mitigation ratios. One could choose the highest ratio of 3 : 1 (aside from infinitely high ratios, i.e. not possible or ‘NP’) so that 3 hectares of restoration are required for every hectare of wetland eliminated by the project. Alternatively, an average or median ratio could be chosen. Assuming project targets are reached within 5 years, additional gains in condition as the wetland reaches ‘maturity’ will offset conditions lost that were below the average. Finally, by acknowledging that functional replacement would be incomplete because function C is not replaceable, even higher ratios could be negotiated, or the project could be completely redesigned. These or other options can be tailored to be responsive to the technical output from rapid assessment. In so doing, the focus is placed on the specifics of how wetlands function and their relative condition, rather than on vague generalities. Instead of statements like ‘the project is damaging to the environment’ and ‘the restoration doesn’t seem adequate’, both the regulator and project proponent are forced to discuss specific reasons why functions are lost, and which options are needed to compensate for the losses. This leads to a related point. The HGM approach provides data that can be used in the

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design of restoration projects through information available on reference wetlands and reference standards. Although not apparent from the superficial treatment given here, data on the variables and indicators from reference standard sites (Table 22.2) can be used for setting project targets and project standards. For a restoration project on forested wetlands to achieve reference standard conditions, more time (i.e. decades) would be required, well beyond the time that most projects are monitored. For forested wetlands, over 50 years may be needed to achieve a relatively mature stand of trees. Even so, the detrital components of such forests will lack the structure normally associated with old growth stands (e.g. VLDW, VDECAY). This is precisely why project standards are needed and must be scrutinised to ensure that they are not in conflict with the ultimate goal of achieving reference standard conditions for restoration projects. The goal is not to restore ecosystems requiring high maintenance, but to launch the mitigation site along a trajectory capable of achieving reference standards. If a restoration site cannot support such levels of functioning because of landscape degradation or other factors, then this constraint should be recognised by acknowledging site potential. The matching of losses and gains becomes more complicated if a temporal perspective is taken. Normally, wetland loss due to project impacts is immediate. Restoration or creation projects require a period of time for wetlands to fully achieve reference standard conditions, especially if restoration begins after the project impact. This creates a ‘deficit’ of functional loss times years. To relieve the deficit, the restored site condition must achieve a level of functioning to match that of the impacted site. This would require higher ratios than those envisioned in the Table 22.5 example. Procedures could be developed to estimate accrual of condition over time. The HGM approach can also be applied at larger scales for determining the functional capacity and condition within a particular geographic area such as a watershed (catchment) planning effort. Gwin et al. (1999) addressed this larger scale in

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the Portland, Oregon metropolitan area by simply classifying wetlands, at the beginning and the end of a 15-year period of urbanisation. During that time, dominance by riverine wetlands (from about 75% of the total) shifted to various types of depression wetlands that were initially extremely rare. Components of the HGM approach are compatible with other approaches that deal with cumulative effects of impacts that alter the condition of wetland ecosystems (Gosselink et al. 1990; Leibowitz et al. 1992). The condition of stream and riparian ecosystems can be estimated at a small watershed scale (10–100 km2) by random sampling of selected reaches for assessment (Rheinhardt et al. 2007).

L I M IT AT IO N S O F T HE HGM AP P R O ACH Any assessment method is no stronger or more reliable than the science information base that underpins it. The reliance of the HGM approach on a reference system provides the backbone for this information, and does so by capturing relevant research, expert interpretation and data on the wetland type from actual sites in the field. The intent of HGM is to provide a platform for this information so that it can be used in the practice of assessing ecosystem change. Given the reality that our knowledge base improves as new information is developed, HGM was designed to be open to new data, alternative interpretations and overall improvement. The modular structure of the HGM approach, which builds upon indicators, variables, model equations of functions and definitions of the functions themselves (Figure 22.6), encourages such change. As such, new information on how wetlands function can be invoked in any of these steps without danger of disarticulating the whole assessment method. The limitations of the approach, as it has been presented in this chapter, are: • it does not provide decision makers with a complete suite of information needed to determine the full consequences of degrading a wetland;

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• only one level of assessment (i.e. rapid) is provided; and the building of the reference system is expensive. In the first instance, the HGM approach does not explicitly identify sites that may have historical significance, endangered species, or other natural or cultural attributes. Other ‘red flag’ features include areas protected as groundwater aquifers, Ramsar sites, wildlife refuges, native lands and so on. (Smith et al. 1995). Further, it does not provide any socio-economic information on impacts, or present results in a format that allows cost-benefit analyses. Each of these would be supplementary to a rapid HGM assessment. Numerous assessments are available, however, to satisfy such needs (Larson and Mazzarese 1994; Fennessy et al. 2004). Because HGM assessments are rapid, they do not provide the detailed information that may be required in a complex or very controversial project. If a project has the potential to release toxic chemicals to adjacent waters, the consequence of such details would not be revealed through a rapid functional assessment. Detailed monitoring and analyses may be required. Further, when an assessment identifies reductions in condition caused by a project and the gains expected in compensatory restoration, the HGM approach makes no policy recommendations on how to handle this output. For example, it is difficult to imagine a project in the real world that would cause loss of one function in the impacted wetland by a given amount, and then provide compensation in the restoration wetland by exactly the same amount and for the same function. Without specific guidance, one can imagine high quality wetlands being lost, to be compensated only with partially restored, degraded wetlands. Policy is also needed if urban wetlands are to be protected, simply because their tendency to be degraded is high, and functional losses from their elimination would be lower relative to many rural wetlands in better condition (Ruhl and Salzman 2006). This and other scenarios clearly indicate the need for a policy framework so the final outcome does not

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The United States HGM Approach result in the ‘tyranny of small decisions’ (Odum 1982). Finally, the work required to build a guidebook (Table 22.4) is substantial. This is necessary to assure, to the extent possible for a rapid assessment, that results will be accurate and repeatable. Consequently, it is not recommended for wetland subclasses that are rare or rarely impacted, nor for resource management programmes that will use them only infrequently. Once developed, however, HGM assessments, as described in this chapter, can be used to screen projects by identifying which functions are likely to be lost by the subclass affected, whether there are any reasonable project alternatives, and if there is potential for compensatory mitigation.

SU M M AR Y AN D CO N CL U S IO NS There are three components of the HGM approach to functional assessment of wetlands: classification by geomorphic setting, articulation of functions and the use of wetland sites to establish reference standards. While each of these is described separately above, in reality, HGM integrates them into one coherent approach. Once the assessment method is developed around standards possessed by the least or minimally altered wetlands of a subclass (the development phase), conducting functional assessments (the application phase) becomes a rapid procedure (usually less than a day) for detecting changes in functioning due to anticipated alterations and restoration projects. Additional characteristics of the approach include the following: The approach should be applicable to ecosystems other than wetlands. The HGM approach to functional assessment is the application of hydrology, geomorphology, community ecology, ecosystem ecology and related disciplines through the use of background information, a logical framework and easily measured ecosystem features.

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The approach does not pretend to be rigorous science. It utilises available scientific information, a logical framework and a platform on which to incorporate professional judgement. In so doing, it is apparent which assumptions underpin the process, what kinds of data are being used and when professional judgement is being applied. The HGM approach identifies information gaps that could be filled with appropriate research. Components of the HGM approach may also be useful as a tool for watershed planning and management. While the approach has been developed for detecting changes in functioning due to impacts and restoration to individual wetlands, it can be modified for other levels of detail. As a screening tool, proposed projects can be provisionally assessed to determine which functions are most likely to be altered. As a cumulative effects procedure, changes in functions and subclasses over time could be monitored at larger scales than routinely used for individual projects. To do so may require the strengthening and incorporation of functions that are expressed at landscape scales. Information is being exchanged with parallel efforts on functional assessment by the European Union (Maltby et al., Chapter 23). The EU is dealing with highly altered landscapes that may confound the use of reference as described in this chapter (Ehrenfeld 2000). Alternatively, the USA effort would benefit from the EU approach for application in urbanising areas. The modular nature of the approach facilitates the incorporation of new information as it becomes available. The models can be regarded as hypotheses, with the expectation that they will be refined as research reveals more about the relationship between ecosystem condition and functions.

ACK N OWLEDGEMEN T S Many people contributed to the development of the HGM approach. Key individuals include

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L.C. Lee, R.D. Smith, D.F. Whigham, E. Maltby, G.G. Hollands, R.D. Rheinhardt, W.L. Nutter, R.P. Novitzki and W.B. Ainslie. This chapter benefited from a review of an earlier version by Ana Inés Malvárez. Manuscript preparation was supported in part by the following: The Center for Transportation and the Environment in cooperation with the US Department of Transportation, the NC Department of Transportation, and the Institute for Transportation Research and Education, North Carolina State University and the National Science Foundation grant BRS87– 02333-04 to the University of Virginia Long Term Ecological Research Program. Figure 22.3 was designed and drafted by Melynda May. Kevin Miller took the photographs for Figure 22.5.

Smith D.R., Ammann A., Bartoldus C. and Brinson, M.M. 1995. An Approach for Assessing Wetland Functions Using Hydrogeomorphic Classification, Reference Wetlands, and Functional Indices. Technical Report WRP-DE-9, US Army Engineer Waterways Experiment Station, Vicksburg, MS. NTIS No. AD A307 121.

National guidebooks National Guidebooks are reviews of concepts and literature as they pertain to general wetland classes, and may serve as templates for development of Regional Guidebooks. They cannot be used to assess a subclass of wetland. Riverine

A P P EN D IX : HGM G U ID E B O O KS This website contains completed guidebooks, related information and literature and archived data: http://el.erdc.usace.army.mil/wetlands/ hgmhp.html. Listed are the guidebooks and other materials developed for the HGM approach, by type of publication and by date. Classification The hydrogeomorphic classification lays a foundation for and support of ongoing efforts to develop methods for assessing wetland condition of physical, chemical and biological functions. Brinson M.M. 1993. A Hydrogeomorphic Classification for Wetlands. Technical Report WRP-DE-4, US Army Engineer Waterways Experiment Station, Vicksburg, MS. NTIS No. AD A270 053.

Approach The approach includes a development and an application phase. The assessment procedure, as the final product, can be used to compare project alternatives, determine impacts, calculate mitigation requirements and so on.

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Brinson M.M., Hauer F.R., Lee L.C., Nutter W.L., Rheinhardt R.D., Smith R.D. and Whigham D. 1995. A Guidebook for Application of Hydrogeomorphic Assessments to Riverine Wetlands. Technical Report WRP-DE-11, US Army Engineer Waterways Experiment Station, Vicksburg, MS. NTIS No. AD A308 365.

Tidal fringe Shafer D.J. and Yozzo D.J. 1998. National Guidebook for Application of Hydrogeomorphic Assessment of Tidal Fringe Wetlands. Technical Report WRPDE-16, US. Army Engineer Waterways Experiment Station, Vicksburg, MS.

Guidelines for developing regional guidebooks Clairain E.J. 2002. Hydrogeomorphic Approach to Assessing Wetland Functions: Guidelines for Developing Regional Guidebooks; Chapter 1, Introduction and Overview of the Hydrogeomorphic Approach. ERDC/EL TR-02-3, US. Army Engineer Research and Development Center, Vicksburg, MS. Smith R.D. 2001. Hydrogeomorphic Approach to Assessing Wetland Functions: Guidelines for Developing Regional Guidebooks – Chapter 3 Developing a Reference Wetland System. ERDC/ EL TR-01-29, US. Army Engineer Research and Development Center, Vicksburg, MS.

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The United States HGM Approach Smith R.D. and Wakeley J.S. 2001. Hydrogeomorphic Approach to Assessing Wetland Functions: Guidelines for Developing Regional Guidebooks – Chapter 4 Developing Assessment Models. ERDC/ EL TR-01-30, US. Army Engineer Research and Development Center, Vicksburg, MS. Wakeley J.S. and Smith R.D. 2001. Hydrogeomorphic Approach to Assessing Wetland Functions: Guidelines for Developing Regional Guidebooks – Chapter 7 Verifying, Field Testing, and Validating Assessment Models. ERDC/EL TR-01-31, US. Army Engineer Research and Development Center, Vicksburg, MS.

Regional guidebooks Ainslie W.B., Smith R.D., Pruitt B.A., Roberts T.H., Sparks E.J., West L., Godshalk G.L. and Miller M.V. 1999. A Regional Guidebook for Assessing the Functions of Low Gradient, Riverine Wetlands in Western Kentucky. Technical Report WRP-DE-17, US. Army Engineer Waterways Experiment Station, Vicksburg, MS. View on-line or download part1.exe & part2.exe. Gilbert M.C., Whited P.M., Clairain E.J., Jr. and Smith R.D. 2006. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Prairie Potholes. ERDC/EL TR-06-5, US. Army Engineer Research and Development Center, Vicksburg, MS. Hauer F.R., Cook B.J., Gilbert M.C., Clairain E.J., Jr. and Smith R.D. 2002. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Intermontane Prairie Pothole Wetlands in the Northern Rocky Mountains. ERDC/EL TR-02-7, US. Army Engineer Research and Development Center, Vicksburg, MS. Hauer F.R., Cook B.J., Gilbert M.C., Clairain E.J., Jr. and Smith R.D. 2002. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Riverine Floodplains in the Northern Rocky Mountains. ERDC/EL TR-02-21, US. Army Engineer Research and Development Center, Vicksburg, MS. Klimas C.V., Murray E.O., Pagan J., Langston H. and Foti T. 2004. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Forested Wetlands in the Delta Region of Arkansas, Lower Mississippi River Alluvial Valley. ERDC/EL TR-04-16, US Army Engineer Research and Development Center, Vicksburg. Download

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Appendix D – Spreadsheets or download Appendix E – Spatial Data. Klimas C.V., Murray E.O., Pagan J., Langston H. and Foti T. 2005. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Forested Wetlands in the West Gulf Coastal Plain Region of Arkansas. ERDC/ EL TR-05-12, US Army Engineer Research and Development Center, Vicksburg, MS. Lin J. 2006. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Depressional Wetlands in the Upper Des Plaines River Basin. Technical Report WRP-DE-17, US Army Engineer Waterways Experiment Station, Vicksburg, MS. Noble C.V., Evans R., McGuire M., Trott K., Davis M. and Clairain E.J., Jr. 2002. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Flats Wetlands in the Everglades. ERDC/EL TR-02–19, US Army Engineer Research and Development Center, Vicksburg, MS. Noble C.V., Evans R., McGuire M., Trott K., Davis M. and Clairain E.J., Jr. 2004. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Depressional Wetlands in Peninsular Florida. ERDC/EL TR-04-3, US Army Engineer Research and Development Center, Vicksburg. Rheinhardt R.D., Rheinhardt M.C. and Brinson M.M. 2002. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Wet Pine Flats on Mineral Soils in the Atlantic and Gulf Coastal Plains. ERDC/EL TR-02-9, US Army Engineer Research and Development Center, Vicksburg, MS. Shafer D.J., Herczeg B., Moulton D.W., Sipocz A., Jaynes K., Rozas L.P., Onuf C.P. and Miller W. 2002. Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Northwest Gulf of Mexico Tidal Fringe Wetlands. ERDC/EL TR-02-5, US Army Engineer Research and Development Center, Vicksburg, MS. Smith R.D. and Klimas C.V. 2002. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Selected Regional Wetland Subclasses, Yazoo Basin, Lower Mississippi River Alluvial Valley. ERDC/EL TR-02-4, US Army Engineer Research and Development Center, Vicksburg, MS. Stutheit R., Gilbert M.C., Whited P.M. and Lawrence K.L. 2004. A Regional Guidebook for Applying the

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Hydrogeomorphic Approach to Assessing Wetland Functions of Rainwater Basin Depressional Wetlands in Nebraska. ERDC/EL TR-04-4, US. Army Engineer Research and Development Center, Vicksburg, MS. Uranowski C., Lin Z., DelCharco M., Huegel C., Garcia J., Bartsch I., Flannery M.S., Miller S.J., Bacheler J. and Ainslie W. 2003. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Low-Gradient, Blackwater Riverine Wetlands in Peninsular Florida. ERDC/EL TR-03-3, US Army Engineer Research and Development Center, Vicksburg, MS. Wilder T.C. and Roberts T.H. 2002. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of LowGradient Riverine Wetlands in Western Tennessee. ERDC/EL TR-02-6, US. Army Engineer Research and Development Center, Vicksburg, MS.

R E F E R E N CE S Adamus P.R. 1983. A Method for Wetland Functional Assessment. FHWA-IP-82-24. Federal Highway Administration, Washington, DC. Adamus P.R., Clairain E.J., Smith R.D. and Young R.E. 1987. Wetland Evaluation Technique (WET) – Volume II. Operational Draft TRY-87. US Army Engineer Waterways Experiment Station, Vicksburg, MS. Ainslie W.B., Smith R.D., Pruitt B.A., Roberts T.H., Sparks E.J., West L., Godshalk G.L. and Miller M.V. 1999. A regional Guidebook for Assessing the Functions of Low Gradient, Riverine Wetlands in Western Kentucky. Technical Report WRP-DE-17, US Army Engineer Waterways Experiment Station, Vicksburg, MS. Armentano T.V. and Menges E.S. 1986. Patterns of change in the carbon balance of organic soil-wetlands of the temperate zone. Journal of Ecology 74, 755–774. Aselmann I. and Crutzen P.J. 1989. Global distribution of natural freshwater wetlands and rice paddies, their net primary productivity, seasonality and possible methane emissions. Journal of Atmospheric Chemistry 8, 307–358. Brinson M.M. 1993a. A Hydrogeomorphic Classification for Wetlands. Technical Report WRP-DE-4. US Army Corps of Engineers Waterways Experiment Station, Vicksburg, MS.

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Brinson M.M. 1993b. Changes in the functioning of wetlands along environmental gradients. Wetlands 13, 65–74. Brinson M.M. 1996. Assessing wetland functions using HGM. National Wetlands Newsletter 18, 10–16. Brinson M.M., Hauer F.R., Lee L.C., Nutter W.L., Rheinhardt R.D., Smith R.D. and Whigham D. 1995. Guidebook for Application of Hydrogeomorphic Assessments to Riverine Wetlands. Technical Report TR-WRP-DE-11, Waterways Experiment Station, Army Corps of Engineers, Vicksburg, MS. Brinson M.M. and Malvárez A.I. 2002. Temperate freshwater wetlands: types, status, and threats. Environmental Conservation 29(2), 115–133. Brinson M.M., Miller K., Rheinhardt R.D., Christian R.R., Meyer G. and O’Neal J. 2006. Developing Reference Data to Identify and Calibrate Indicators of Riparian Ecosystem Condition in Rural Coastal Plain Landscapes in North Carolina. Report to the Ecosystem Enhancement Program, North Carolina. Department of Environment and Natural Resources, Raleigh, NC, 62 pp. Brinson M.M. and Rheinhardt R. 1996. The role of reference wetlands in functional assessment and mitigation. Ecological Applications 6, 69–76. Brinson M.M. and Rheinhardt R.D. 1998. Chapter 2. Wetland functions and relations to societal values. In: Messina M.G. and Conner W.H. (editors), Southern Forested Wetlands: Ecology and Management, pp. 29–48. Lewis Publishers, Boca Raton, FL. Ehrenfeld J.G. 2000. Evaluating wetlands within an urban context. Ecological Engineering 15, 253–265. Federal Register 1997. National action plan to implement the hydrogeomorphic approach to assessing wetland functions. Federal Register 62(119), 33607–33620. Fennessy M.S., Jacobs A.D. and Kentula M.E. 2004. Review of Rapid Methods for Assessing Wetland Condition. EPA/620/R-04/009. US. Environmental Protection Agency, Washington, DC. Golet R.C, Calhoun A.J.K., DeRagon W.R., Lowry D.J. and Gold A.J. 1993. Ecology of Red Maple Swamps in the Glaciated Northeast: A Community Profile. US Department of Interior, Fish and Wildlife Service, Biological Report 12, Washington, DC. Gosselink J.G., Shaffer G.P., Lee L.C., Burdick D.M., Childers D.L., Leibowitz N.C., Hamilton S.C., Boumans R., Cushman D., Fields S., et al. 1990. Landscape conservation in a forested wetland watershed. BioScience 40, 588–600.

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The United States HGM Approach Gwin S.E., Kentula M.E. and Shaffer P.W. 1999. Evaluating the effects of wetland regulation through hydrogeomorphic classification and landscape profiles. Wetlands 19, 477–789. Larson J.S. and Mazzarese D.B. 1994. Rapid assessment of wetlands: history and application to management. In: Mitsch W.J. (editor), Global Wetlands: Old World and New. Elsevier, Amsterdam, pp. 625–636. Leibowitz S.G., Abbruzzese B., Adamus P.R., Hughes L.E. and Irish J.T. 1992. A Synoptic Approach to Cumulative Impact Assessment: A Proposed Methodology. EPA/600/R-92/167. US Environmental Protection Agency, Environmental Research Laboratory, Corvallis, OR. Lugo A.E. and Snedaker S.C. 1974. The ecology of mangroves. Annual Review of Ecology and Systematics 5, 39–64. Magee T.K., Ernst T.L., Kentula M.E. and Dwire D.A. 1999. Floristic comparison of freshwater wetlands in an urbanising environment. Wetlands 19, 517–534. Maltby E., Hogan D.V., Immirzi C.P., Tellam J.H. and van der Peijl M.J. 1994. Building a new approach to the investigation and assessment of wetland ecosystem functioning. In: Mitsch W.J. (editor), Global Wetlands: Old World and New. Elsevier, Amsterdam, pp 637–658. Maltby E., Hogan D.V. and McInnes R.J. (editors) 1996. Functional Analysis of European Wetland Ecosystems, Phase 1. Ecosystems Research Report No 18, Final Report EC DG XII CT90–0084. Office for Official Publications of the European Communities, Luxembourg, Brussels. McCune B. and Mefford M.J. 1995. PC-ORD. Multivariate Analysis of Ecological Data, Version 3.0. MjM Software Design, Gleneden Beach, OR. NRC (National Research Council). 2002. Riparian Areas: Functions and Strategies for Management. National Academy Press, Washington, DC, 386 pp. Odum W.E. 1982. Environmental degradation and the tyranny of small decisions. BioScience 32, 728–729. Odum W.E., Smith III T.J., Hoover J.K. and McIvor C.C. 1984. The Ecology of Tidal Freshwater Marshes of the United States East Coast: a Community Profile. FWS/OBS-83/17, Fish and Wildlife Service, Washington, DC. Rheinhardt R.D., Brinson M.M., Christian R.R., Miller K.H. and Meyer G.F. 2007. A reference-based framework for evaluating the ecological condition of stream networks in small watersheds. Wetlands 27(3), 524–542.

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Rheinhardt R.D., Rheinhardt M.C., Brinson M.M. and Faser K.E. Jr. 1999. Application of reference data for assessing and restoring headwater ecosystems. Ecological Restoration 7(3), 241–251. Rheinhardt R.D., Rheinhardt M.C. and Brinson M.M. 2002. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Wet Pine Flats on Mineral Soils in the Atlantic and Gulf Coastal Plains. Report ERDC/EL TR-02-9, US. Army Corps of Engineers, Engineer Research and Development Center, Vicksburg, MS. Rhul J.B. and Salzman J. 2006. The effects of wetland mitigation banking on people. National Wetlands Newsletter 24(1), 9–14. Shaffer P.W. and Ernst T.L. 1999. Distribution of organic matter in freshwater emergent/open water wetlands in the Portland, Oregon metropolitan area. Wetlands 19, 505–516. Shaffer P.W., Kentula M.E. and Gwin S.E. 1999. Characterisation of wetland hydrology using hydrogeomorphic classification. Wetlands 19, 490–504. Smith R.D., Ammann A., Bartoldus C. and Brinson M.M. 1995. An Approach for Assessing Wetland Functions using Hydrogeomorphic Classification, Reference Wetlands, and Functional Indices. US Army Corps of Engineers Waterways Experiment Station, Technical Report TR WRP-DE-10. Vicksburg, MS. State of Alaska Department of Environmental Conservation/US Army Corps of Engineers Waterways Experiment Station Technical Report Number WRP-DE. 1999. Operational Draft Guidebook for Reference Based Assessment of the Functions of Precipitation-Driven Wetlands on Discontinuous Permafrost in Interior Alaska. Anchorage, AK, available at http://www.dec.state.ak.us/water/wnpspc/ wetlands/interior_operational_draft_may_1999b.pdf, last accessed on 28 January 2009. Stutheit R., Gilbert M.C., Whited P.M. and Lawrence K.L. 2004. A Regional Guidebook for Applying the Hydrogeomorphic Approach to Assessing Wetland Functions of Rainwater Basin Depressional Wetlands in Nebraska. ERDC/EL TR-04-4, US. Army Engineer Research and Development Center, Vicksburg, MS, available at http://el.erdc.usace.army.mil/wetlands/ pdfs/trel04-4.pdf, last accessed on 28 January 2009. Trimble S.W. 1970. The Alcovy River swamps: The result of culturally accelerated sedimentation. Bulletin of the Georgia Academy of Science 28, 131–144. US Fish and Wildlife Service. 1981. Habitat Evaluation Procedures. FWS/DES-ESM 102, Washington, DC.

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Wharton C.H., Kitchens W.M., Pendleton E.C. and Sipe T.W. 1982. The Ecology of Bottomland Hardwood Swamps of the Southeast: A Community Profile. US. Fish and Wildlife Service, Biological Services Program, Washington, DC, FWS/OBS-81/37. Whigham D.F., Chitterling C. and Palmer B. 1988. Impacts of freshwater wetlands on water quality; a landscape perspective. Environmental Management 12, 663–671.

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Whigham D.F., Lee L.C., Brinson M.M., Rheinhardt R.D., Rains M.C., Mason J.A., Kahn H., Ruhlman M.B. and Nutter W.L. 1999. Hydrogeomorphic (HGM) assessment – a test of user consistency. Wetlands 19(3), 560–569. Wofsy S.C. 2001. Climate change: Where has all of the carbon gone? Science 292, 2261–2263.

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23 Development of a European Methodology for the Functional Assessment of Wetlands E D W AR D MA L T B Y , T OM BARKER AND CO NO R LI NSTEAD Institute for Sustainable Water, Integrated Management and Ecosystem Research, University of Liverpool, Liverpool, UK

IN T R O D U CT ION Managers of wetlands have often been content to maintain overall ecosystem character through structure, species composition and habitat appearance, but how much does a manager know about how the wetland is actually working? Does its functioning deliver potential benefits, such as floodwater retention, groundwater recharge, removal of contaminants and maintenance of high value biodiversity? These benefits are often referred to as ‘ecosystem services’ (Ehrlich and Mooney 1983; MA 2005a). If a wetland suffers from environmental impacts, can management be changed to enhance the particular processes necessary to restore specific aspects of functioning? One of the key objectives of wetland science is ‘the development of refined wetland classification and evaluation systems; improved tools … for better wetland management; and education of government officials, land owners, engineers, lawyers and biologists on wetland functions, values and techniques for management’ (Mitsch 1994). This is one of the most demanding challenges for the science research community. If significant wetland resources are to be maintained

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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for the future, their importance beyond solely academic interest must be recognised by society. Reliable ways must be developed to assess the functions that deliver benefits valued by society. This requires transformation of available technical knowledge into practical tools for environmental management and conservation. In this chapter, we discuss key issues in the translation of empirical studies and expert knowledge into robust, practical procedures for use in assessing wetland functioning. Assessment of functioning is a key prerequisite to making management decisions that affect delivery, by wetlands, of specific or particular combinations of ecosystem services. This knowledge is critical to provision of the evidence necessary to underpin strategic and policy decisions. Effective functional assessment provides an essential tool for those individuals, regulatory bodies, and other government or nongovernmental organisations that make informed decisions for appropriate wetland management. The dominant policy approach to wetland protection and management in Europe has been based largely on traditional nature conservation criteria, such as biodiversity, naturalness, rarity and typicality (Maltby et al. 1994). The success of this approach has relied on the formal designation of specific wetland sites as targets for arresting loss, and on the assumption that such status provides a barrier, or at least a deterrent,

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to adverse external impacts. Such measures of conservation value have been implemented generally through various tiers of ‘protected areas’ supported by appropriate legislation. Whilst through this approach many wetlands have been identified and recognised as intrinsically important, the strategy is generally insufficiently relevant to the wider interests of society. A case in point is the recent intervention by the European Commission to stop, possibly permanently, a major road project at Rospuda, Poland, one of the continent’s last remaining pristine fens, after the Polish government offered ‘weak and unconvincing’ mitigation measures (Stigler 2007). There are many other examples where wetlands have been lost or degraded by human activities, both deliberate and unintentional, despite the apparent strength of ‘protective’ conservation legislation. A prime example is that of the wetlands of the Tablas de Daimiel National Park, Spain, which were severely degraded by groundwater abstraction from the surrounding cultivated area via the La Mancha aquifer in the 1970s. Extreme desiccation combined with high temperatures eventually led to combustion of the peat substrate, and extensive fire damage to the ‘protected’ ecosystem. Emergency remediation measures have included dam-building, well-drilling and transfers of water, in attempts to avert permanent loss of the wetlands (Llamas et al. 1996). Although a necessarily high conservation profile is given to wetland ‘jewels’, this is not in itself a guarantee of conservation or appropriate management. The limitations of conservation protection are only too apparent. For example, in the Lower Guadalquivir, Spain, public pressure in the 1960s demanded protection of the Doñana wetlands, world renowned for birdlife, after they were threatened with drainage and development, but in the 1990s the area was again under threat from proposals to build a resort, the Costa Doñana, at its edge. Further public pressure led to designation of the Doñana Nature Park, covering 100 000 hectares. Such designation, however, still fails to protect the underlying aquifer, which supports the rich wetland complex from over-

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exploitation for irrigation and other uses outside the park (Llamas 1992; UNEP-WCMC 2007). Whilst the conservation approach has had many successes, it has tended to polarise an apparent ‘soft’ nature conservation ethic against the perceived immediate socio-economic needs of society. It also fails to highlight the importance of the much greater wetland resource that will always exist alongside any formal protected area network. Perhaps more than any other group of ecosystems, protected wetland areas are vulnerable to impacts arising outside their boundaries. This is because of the frequent dependence on external water flow or aquifers that are themselves subject to environmental pressures over extensive areas (see Baker et al., Chapter 6; Gilvear and Bradley, Chapter 7; Grootjans and Van Diggelen, Chapter 8; McCartney and Acreman Chapter 17). Ecosystem management that is restricted to areas within the wetland boundary may be insufficient to secure desired environmental conditions in the wetland (Hollis et al. 1988). The limitations faced by a strategy of habitatbased wetland protection include scale, resource and connectivity issues. Many wetlands in Europe occur in small landscape fragments (though with some notable exceptions) and often are important parts of traditional rural economies. This situation presents a challenge to conservation groups, such as those concerned with biodiversity management, and to other stakeholders who may have potentially conflicting objectives such as cultivation or livestock grazing. Additionally, only limited resources are available to bring wetlands within the protected area network, hence only a small proportion of the total wetland resource can ever be managed in this way. Even international protection under the Ramsar Convention, dedicated exclusively to wetlands, has not proven sufficient to protect valuable wetlands in the face of powerful commercial and political interests. Government complicity was suspected when Chile’s first Ramsar site, the Carlos Anwandter Nature Sanctuary in Rio Cruces, suffered pollution from an economically important mill (Marcotte 2006). The wetland, home to many species of animal and plant,

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Functional Assessment of Wetlands including a large and important population of black-necked swans (Cygnus melancoryphus), was subject to heavy contamination from an upstream paper mill that failed to comply with official limits to the amount and type of discharges. Serious ecological consequences included the loss of the swans (Mulsow and Grandjean 2006). Whatever the level of legal protection a wetland may have, unless it is possible to manage activities across the wider catchment and other areas contributing to the functioning of the wetland, it will always be difficult to avoid external effects that can degrade the ecosystem. This requirement is now implicitly recognised in the catchment-based approach to water resource management in the European Water Framework Directive (2000/60/EC). In the same way that wetlands are susceptible to events and pressures in their catchment, they may also influence environmental conditions downstream, and this can produce considerable benefits to society and the local environment. Such benefits include flood protection, delivery of clean water and support for migratory fisheries. Wetland losses have been allowed to occur because the full significance and value of the work done by intact and healthy wetland ecosystems goes largely unmeasured and unappreciated. It is this deficiency in the recognition of the services provided by natural ecosystem functioning, and the rectification of it, that underpins the scientific developments examined in this chapter and their application in wetland ecosystems. Importance of services The tangible benefits wetlands bring to the wider environment are achieved through a multiplicity of interactions within the ecosystem. Nutrients and particulates in inflowing water or from the atmosphere are captured in the physical structure of the wetland. As algae and plants take up nutrients from the water, soil and sediments, and incorporate them as biomass, a wide range of heterotrophic organisms, ranging from bacteria and fungi crustaceans, worms, beetles, flies and their myriad larvae, to higher animals such as

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fish, amphibians, birds, reptiles and mammals, including humans, variously process, remove and add nutrients, contaminants and organic matter. The culmination of many biological and chemical processes characterises the particular pattern of functioning of the wetland, which can result in an impressive range of ecosystem services that have regulatory, resource and social values. The timber, wild foods and freshwater produced in wetlands are just some of the most obvious products. Wetlands support the provision of a wide range of goods and services useful to humans (Figure 23.1). The multiple values of wetlands, however, are not always recognised. Decisions that have led to the demise of wetlands rarely, if ever, have considered their importance as functioning ecosystems relevant both to wildlife and the quality of human life. In a 4-year study, the Millennium Ecosystem Assessment (MA) attributed freshwater wetland losses to direct loss of habitat, over-exploitation of resources, impacts from invasive species, pollution by nutrient (nitrogen and phosphorus) runoff and climate change, all of which are expected to remain important drivers of wetland loss (MA 2005a). A range of factors have facilitated the degradation and loss of wetlands. In particular these include industrial expansion and unsustainable levels of exploitation of natural resources. Generally, the significance of wetlands has been appreciated only after their alteration or loss, and usually when it is too late or expensive to rectify the damage. Whilst it is often more readily acknowledged that wetlands can present hazards to animal and human health by hosting vectors of disease, government conservation agencies have generally failed to recognise the specific roles of wetlands in the provision of goods and services for direct or indirect human use, the maintenance or enhancement of wider environmental quality, including human health, and their role in the achievement of sustainable development objectives. This view, however, is gradually changing. In England and Wales the Department of the Environment, Food and Rural Affairs (Defra) has recently initiated a programme to establish the evidence base for the delivery of ecosystem services, but there is a long way to go before any formal recognition of

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Wetland ecosystems Structure Geomorphology Hydrology Soils Biota

Interactions

Processes Physical Chemical Biological

Wetland functioning Hydrological functions Biogeochemical functions Ecological functions

Ecosystem services Supporting

Provisioning

Regulating

Cultural

• Nutrient cycling • Primary production • Secondary production • Soil-building

• Food • Fodder • Fresh water • Wood • Fibre • Fuel • Genes

• Climate • Floods • Storms • Diseases • Water • Toxins • Pollution

• Aesthetics • Religion • Education • Recreation • Tourism

Fig. 23.1 Ecosystem services are the ultimate products of physical, chemical and biological interactions operating within the wetland. Collectively, these perform the functions that provide both tangible and intangible ecosystem services supporting environmental and socio-economic security.

ecosystem services becomes central to key sector policies (Defra 2007a,b). Wetland functioning The physical, chemical and biological properties, and the processes that occur within wetland ecosystems, individually or in combination, control functioning, which in turn results in the provision of valuable goods and services (Figure 23.1; see also Maltby, Chapter 1 (introduction); Roggeri, Chapter 25; Turner et al., Chapter 26; Ross and Murkin, Chapter 35; and all of Section 7). Recognition of the importance of ecosystem services underpins the concept of natural capital (Costanza et al. 1997) and is pivotal in the MA,

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which focuses on connections between ecosystem services and the values they provide (MA 2005a). The MA categorises ecosystem services to include provision (food, freshwater, wood and fibre products, and fuel), regulation (of climate, flooding, and disease and purification of water), and cultural benefits (aesthetic, spiritual, educational and recreational), all underpinned by the supporting processes enabling nutrient cycling, soil formation and primary production. Dependent on these are the ‘constituents of wellbeing’ that include security (personal safety, secure resource access and security from disasters), the basic materials for good life (adequate livelihoods, sufficient nutritious food, shelter and access to goods), health (strength, feeling well and access to clean air and water), and good

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Functional Assessment of Wetlands social relations (social cohesion, mutual respect and the ability to help others). Collectively these give people freedom of choice and action, defined as the ‘opportunity to be able to achieve what an individual values doing and being’ (MA 2005b). The MA, and increasingly accessible scientific knowledge, have contributed to a growing appreciation of the benefits to society of services provided by ecosystems. The increasing consequences of ignoring the roles of wetlands in delivering benefits to both the natural and socio-economic landscapes have given an urgent political dimension to wetland management in its role as a vital element in the attainment of sustainable development. Social and economic aspects of wetland functioning, through the delivery of ecosystem services, may be of greater concern to the general public than specific considerations of biodiversity. There is likely to be public support for wetland conservation where economic values can be identified, for example where, as part of their normal functioning, wetlands reduce flooding hazard, reduce nutrient loads or filter out contaminants, thus providing valuable services such as public safety from floods, clean water for drinking, healthy fisheries for harvest or recreation and pollinators for agriculture. A widened policy approach to wetlands to include the broader functional dimension takes us: • from an emphasis on individual sites to consideration of the greater wetland resource in the landscape; • from a simple sectoral view to a multi-sector perspective; and • to a framework that highlights the role of wetlands in contributing to basic human needs and welfare as well as underpinning environmental and economic sustainability. A functional view of wetlands There is an obvious need for the development of methods that can identify, and preferably quantify, the ways in which wetlands work, and can relate this knowledge to the overall socio-economic

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benefits to society achievable from the conservation and management of wetlands. The adoption of a functional approach to wetland evaluation effectively broadens the political significance of wetlands from icons of conservation, to the complex and functionally-varied providers of ecosystem services that they are. Crucially, decision-makers must be given sufficient information to make objective assessments of individual functions or groups of functions that underpin the delivery of ecosystem services. From this, they can make more evidence-based decisions aligned with sectoral objectives or offer alternatives to stakeholders, enabling better informed debate and choices. There are at least three significant challenges to the development of a functional approach: i Not all wetlands perform all functions, nor do they perform individual functions to the same extent (Table 23.1); ii It would be prohibitively expensive, and impossible practically, to determine or verify the predicted functioning of each and every wetland by means of empirical research or monitoring; iii There is still considerable debate over exactly how the economic values arising from the functioning of wetlands (and indeed other types of ecosystem) should be derived and expressed (Turner et al., Chapter 26). Key operational goals for the use of functional assessments are: • to facilitate better wetland management decisions, based on good science and expert judgement, for the long-term benefit of society; • to avoid the constraints of time, expense and limited expertise that often face empirical research; • to prevent gross misinterpretations of wetland functioning due to generalisations; • to recognise that different parts of a single wetland may function in different ways; • to provide research-based support for the targets and priorities of practical wetland restoration; and • to translate fundamental scientific understanding into transferable knowledge that can be used by decision-makers.

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In addition to gaseous removal of N, nutrients may also be exported at varying rates from the wetland owing to remobilisation and delayed transportation.

Carbon may be stored for long periods, often as peat, which can be useful in mitigation of global warming. Peat supports rare and valuable biological communities.

In-situ C retention

Sediment retention

Sediment retention

Nutrient export

Groundwater discharge

Groundwater discharge

Retaining or delaying the release of the nutrients N and P into water bodies may prevent potentially deleterious effects on watercourses. Eutrophication changes the ecosystem composition and character, and adversely affects potable water quality.

Groundwater recharge

Groundwater recharge

Decreases peak flow in rivers, reducing flood damage downstream. Provides wildlife habitat especially important for fisheries support. Replenishment of groundwater resources. Maintenance of dependent ecosystems in discharge areas. Maintenance of base flow in rivers fed by discharge elsewhere. Emergence at springs or seepage zones. Maintenance of river base flow. Maintenance of ecosystems dependent on soil–water regimes. Improvement of river water quality due to the reduced input of suspended sediments and associated sediment nutrients.

Service provision

Nutrient retention

Floodwater detention

Hydrological: Floodwater retention

Biogeochemical: Long-term retention of nutrients (N and P) through plant uptake. Storage of nutrients (N and P) in soil organic matter. Adsorption of N as ammonium. Adsorption and precipitation of P in the soil. Retention of particulate nutrients (N and P). Gaseous export of N: Denitrification; Ammonia volatilisation. Export of nutrients through vegetation management. Export of nutrients via wind and water-mediated processes. Organic matter accumulation.

Wetland function

Wetland process

Table 23.1 A range of wetland processes, functions and services normally provided by ‘healthy’ wetland ecosystems. The provision of ecosystem services depends on wetland functioning, which is itself dependent on the operation of natural wetland processes. (From Maltby 2009.)

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The physical, chemical, and biological processes occurring in wetlands all contribute to the provision of unique habitats, which support a variety of adapted organisms. This contributes to global biodiversity, and supports recreation and quality of life.

The food web (network of organisms based on intra- and inter-specific energy transfers) of the wetland may be supported by biomass production on-site or by detritus and organisms transported into the system. The production of biomass by the wetland can also be responsible for supporting food webs at other sites.

Food web support

Organic carbon concentration control

Organic carbon export into surface waters.

Improvement of river water quality due to the reduction of suspended or dissolved trace element loads. Trace element loads can have toxic effects on the stream ecosystem. Prevention of groundwater contamination and uncontrolled translocation of trace elements within the river marginal wetland. When soil or sediment storage capacity is exceeded, there can be consequent risks for food production (through remobilisation by plant uptake). Possibility of controlled removal of trace elements, but danger of toxic effects on plants, and recontamination of river water or groundwater. Wetlands strongly influence the concentration of dissolved organic carbon in runoff water, key features of water quality and the aquatic ecosystem in areas with DOC-rich water (over c. 5 mg L−1 DOC).

Ecosystem maintenance

Trace element export

Plant uptake of trace elements. Physical remobilisation of trace elements. Biogeochemical remobilisation of trace elements.

Ecological: Provision of overall habitat structural diversity. Provision of microsites for: macroinvertebrates fish herptiles birds mammals Provision of plant and habitat diversity. Biomass production. Biomass import via physical processes: watercourses overland flow wind transport Biomass import via biological processes: fauna; anthropogenic means; Biomass export via physical processes Biomass export by biological processes

Trace element storage

Physical retention of trace elements. Biogeochemical retention of trace elements.

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In North America, methods for the assessment of wetland functioning have been high on the research and policy agendas for several decades, and a multitude of methodologies have been produced to meet a range of operational requirements (see Larson, Chapter 21; Brinson, Chapter 22; Smith, Chapter 24). A review by Bartoldus (1999) has characterised some 40 different wetland assessment procedures for the United States alone. The diverse methodologies reflect the rich variety of habitats, species addressed, time needed, costs, scales, outputs, expertise required, uses and potential users. Within the United States, the purpose of wetland assessment has been defined as the means to better inform decision-makers of the public value of wetland functions that may be lost or impaired by development projects (Larson and Mazzarese 1994; Larson, Chapter 21). Almost inevitably, legal disputes have arisen involving the rights of individuals and proposed activities in wetlands, and has led to the concept of ‘jurisdictional wetlands’, with an emphasis on the need to delineate wetland areas on the basis of scientifically rigorous as well as legally verifiable criteria. Attempts to develop tools for the rapid assessment of wetland functioning have tended to treat a wetland site as a single functional unit. Pioneering work on assessing wetland functions and values in the United States (e.g. Adamus 1983; Lonard and Clairain 1985) led to the development of the Wetland Evaluation Technique (WET) (Adamus et al. 1987), which was based fundamentally on the assessment of an entire wetland area or multiple ‘assessment areas’, representing spatially distinct parts of larger wetlands. The simple rating outputs of the system were considered inadequate for decision-making (Adamus et al. 1991). In the 1990s, Smith et al. (1995) presented an approach based on a hydrogeomorphic (HGM) classification of wetlands, and Brinson (1996) elaborated the advantages of the HGM approach based on a standard, hydrologically-biased classification, the use of ‘reference’ wetlands and the distinction between wetland functions and values, often previously

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grouped together (see Brinson, Chapter 22). Partly because of the limited extent of precise scientific information about process variation at smaller scales, and partly because of the limited need for application at any other scale, the HGM method is based on the assessment of entire wetlands, rather than subdivisions with spatial and temporal variations in functional behaviour. There is a notable absence of similar wetland assessment methods in Europe, yet the need is no less important; a fact highlighted by a combination of policy statements. These include the 1995 European Commission Communication to the Council of the European Parliament on the ‘wise use and conservation of wetlands’, which recognised the critical condition of Europe’s wetland resources and the need for urgent action, and the Water Framework Directive, which provides the opportunity to achieve conservation of important wetlands, as well as making progress towards the goal of attaining good ecological condition of European water bodies, partly through the maintenance of wetland functioning (EU 2003). The differences between European and American wetlands mean that the United States’ and Canadian approaches are generally unsuitable for direct application in Europe. The main reasons for this are: • the prominence of management in European wetlands, and the importance of their connections with agriculture and rural practices; • the small size of the majority of European wetlands, their extreme regional and geographical diversity and, in many cases, their intimate role in the culture and socio-economics of local communities; • the limited attention historically given to wetlands as a distinctive area of scientific enquiry, and the consequent inadequacy of a consolidated, transdisciplinary science base (Maltby et al. 1994; Bullock and Acreman 2003); • the lack of a strong policy or regulatory framework specifically for wetlands in Europe (cf. Section 404 of the Clean Water Act in the United States; see Smith, Chapter 24). The new Water Framework Directive (WFD) is the nearest Europe has to a broad policy for the protection of wetlands. It requires the

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Functional Assessment of Wetlands establishment of integrated river basin management plans to achieve ‘good ecological status’ of river, lake, estuary and coastal water bodies by 2015. The WFD considers wetlands to be distinct from other water bodies, and not its primary focus. The European Commission has published guidance on the role of wetlands in the WFD under its Common Implementation Strategy (EU 2003) but this has yet to have any significant influence on wetland resource management at either the site or catchment scale.

Developing functional analysis for application to European wetland ecosystems The development of a means to assess wetland functioning in the European environment has been based on a long-term interdisciplinary panEuropean science research effort funded substantially by the European Commission (Table 23.2 indicates the different initiatives). This work has culminated in the development of functional assessment procedures, commonly referred to as

Table 23.2 Sites investigated in the development of the functional assessment procedures. Further information on these projects can be found by consulting CORDIS: http://cordis.europa.eu/ Name and location

Project acronym

Description

Purpose of study

River Torridge/Walden, UK

FAEWE

Nutrient enrichment impact

River Shannon/Little Brosna, Ireland River Loire/Allier, France

FAEWE

River Guadiana headwaters, Spain River Danube, Romania

FAEWE

Lac de Grande Lieu, France

PROTOWET

Tour du Vala, France

PROTOWET

River Svartberget headwaters, Sweden Oostvaarderplassen, The Netherlands River Mulde/Elbe, Germany

PROTOWET

River marginal wetlands (wet grassland) on floodplains and footslopes River marginal wetlands (wet grassland – callows) on floodplains River marginal wetlands (dry and wet grassland, scrub) on floodplains River marginal wetlands (sedge and reed beds and salt marsh) River marginal wetlands (floodplain grassland and forest) and polders (rice) Lake marginal wetlands (reed beds and humid grassland) Swamps, wet grassland and lakes of the Camargue River marginal wetlands (coniferous forest)

PROTOWET

Lake marginal wetlands (reed beds)

PROTOWET

River Tamar/Carey, UK (Tamar catchment)

EVALUWET

Waterland catchment, The Netherlands Hovran catchment, Sweden Paríž Creek catchment Neajlov catchment, Romania Lake Cheimaditida catchment, Greece Elbe catchment, Germany

EVALUWET

River marginal wetlands (floodplain grassland, reed and sedge beds) River marginal wetlands (floodplain wet grassland and footslope seepage zones of wet grassland and woodland) Fen meadows, lakes

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FAEWE

FAEWE II

EVALUWET EVALUWET EVALUWET EVALUWET EVALUWET

Wet pasture, forests, and lakes Reed beds, wet grassland Marshes, lakes, and artificial ponds Lake marginal reed beds, wet meadows, and calcareous fens Floodplain hardwood forest, floodplain wet grassland, oxbow lakes with marginal swamps

Sediment deposition impact River flow regulation impact Groundwater abstraction impact Flooding regimes and polders impacts Grazing impact Grazing impact Afforestation and drainage impacts Grazing and nutrient enrichment impacts Heavy metals impact FAP testing

FAP testing FAP testing FAP testing FAP testing FAP testing FAP testing

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FAPs (Maltby 2009) and their incorporation into a versatile GIS-based Wetland Ecosystem Decision Support System (WEDSS 2007).

A C O N CE P T U AL F R AM E W OR K The character of hydrological, biogeochemical and ecological processes within a wetland, their rates and their contributions to functioning, are the result of definable site characteristics, which themselves may vary both within the wetland and in its catchment. The FAPs are built on the identification of relationships between wetland ecosystem processes and observable site properties, resulting in the definition of more or less distinctive landscape subunits with particular functional characteristics. The approach assumes that it is possible: • to recognise wetland units in the landscape that are likely to function in particular and predictable ways; • to identify such units on the basis of easyto-use criteria; • to infer processes from observable properties or indicators; and • to assess functioning through identification of particular processes individually or in combination. Consultations with potential users indicated that the specifications of an assessment procedure should enable application by those who may not necessarily have in-depth knowledge of wetland ecosystems or their functioning, should be flexible enough for implementation by a variety of users with wide-ranging interests, and be capable of relatively rapid deployment (on a time scale of hours to days rather than weeks to months). The detailed structure of the FAPs was developed in consultation with potential users of the system across Europe, notably in the UK and the Netherlands, where consultative groups were formed. These consultations established that an effective assessment procedure should: • assist planners in their decision-making for development and catchment management planning; • identify levels of environmental impact that alter functioning;

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• provide a scientific basis for the protection and appropriate management of wetlands; • give guidance on the management necessary to optimise wetland functioning; • where appropriate, enable quantification or semi-quantification of functioning; • provide guidance to assist non-experts to conduct assessments; and • help identify potential targets for restoration and the means to achieving them. The FAPs have been designed to meet these requirements. For decision-makers, the method offers understanding of wetland functioning. The assessments can supplement other information relating to societal priorities, costs and policy limitations, in order to assist and support relevant decision-making. The FAPs translate up-to-date scientific knowledge into reasonable predictions of how different parts of wetlands function in different landscapes. The complexity of individual wetland ecosystems is accommodated through identification of functionally distinct hydrogeomorphic units (HGMUs), conceptualised in Maltby et al. (1994). The FAPs outcomes are linked to socially relevant priorities including flood control, water quality and biodiversity conservation. The conceptual framework for development of the functional assessment is based on a simple process-response model (Figure 23.2) explained in detail elsewhere (Maltby et al. 1994, 1996). The Hydrogeomorphic Unit (HGMU) There are at least three reasons why functioning is unlikely to be uniform throughout the area of a wetland. First, hydrology commonly exhibits high spatial variability within individual wetlands related, for example, to subtle differences in elevation or to a different balance of surface water and groundwater over a site. Second, wetlands themselves frequently are ‘ecotones’ at the interface between terrestrial and aquatic ecosystems and, accordingly, exhibit more or less continuous variation in functioning across the transition between dry and wet conditions. Third, natural

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Natural environmental characteristics e.g. hydrological regime geomorphological context geology climate

Wetland process domain food chain dynamics ion dynamics succession/stress water routing

Value 1 e.g. water quality

Natural wetland ecosystem N1

Elements Properties Interactions

Value 2 e.g. habitat

Ecosystem functioning

Value 3 e.g. food chain support Environmental change: Temperature Rainfall Runoff Groundwater Contaminants

Direct anthropogenic impacts: e.g. drainage flooding regime fertilisers grazing Altered: elements properties interactions

Ecosystem variables: Water table Hydrological regime Nutrient dynamics Plant/animal communities Threshold 1 Altered wetland ecosystem W1

Increasing stress level

Modified process domain Threshold 2

W2

Threshold 3 W3

Fig. 23.2 A simple process-response model of the assessment for the evaluation of wetland functioning under different environmental stresses that alter physical, chemical and biological processes. (Based on Maltby et al. 1994.)

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wetland processes are often locally altered or disrupted by land management or other human impacts, for example drainage, grazing and fertiliser applications. The FAPs divide a site into areas considered to be the smallest practical identifiable units in the wetland landscape that exhibit functional homogeneity. These functional units are designated on the basis of hydrogeomorphic criteria, where geomorphology is described in terms of slope (gradient), depressions and elevations. Hydrology is assessed according to differences in surface and near-surface inflows, outflows, and management. Additionally, soil is used as a long-term proxy indicator of hydrogeomorphic conditions. The term hydrogeomorphic unit (HGMU) has been adopted to describe the distinct, hydrogeomorphologically-defined areas. The HGMU can be defined formally as ‘a landscape unit of uniform geomorphology and hydrological regime, where soils are also uniform in that they are a reflection of the hydrology and geomorphology’ (this is slightly modified from the original definition in Maltby et al. 1996). Vegetation is not used as a defining characteristic because of its dependence on land use and management. Instead it is described for each HGMU, enabling inference of important information on the hydrology or trophic status of the unit. Hydrogeomorphic unit designation provides the physical template for assessment in a comparable way to that used at an overall site scale in the United States (Brinson et al. 1994; Brinson, Chapter 22), but hydrological and geomorphological controls are also considered at larger landscape scales pertinent to the site (Semeniuk and Semeniuk 1995). Whilst the US model uses hydrogeomorphology at the siteto-landscape scale, initially at least, the FAPs have applied the principle to the subsite scale. More recent applications, however, have been exploring the ways in which the procedures can be scaled up to larger landscape HGMUs to facilitate broader management strategies, for example in the Humberhead Levels, UK (below).

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Establishing a science base The FAPs are underpinned by extensive wetland research, undertaken at representative sites along European environmental gradients. Four study areas were selected initially, along a climatic gradient from the humid conditions of Ireland, through south-west England to the continental regime of central France and the highly seasonal, semi-arid regime of central Spain. Paired sites were selected to measure the relative effects of specific anthropogenic impacts (sedimentation, nutrient enrichment, river-flow regulation and groundwater abstraction). Detailed physical characteristics, together with a range of process studies (Table 23.3), provided the scope for the study sites to act as calibration points along the wider spectrum of European wetland ecosystems (Maltby et al. 1996). This supported recommendations by Brinson (1991) in the United States, who has argued consistently that the process of wetland assessment ‘could benefit greatly by the establishment of a nation-wide network of “calibration wetlands” as focal points of research into better assessment indicators and as sites for training personnel that conduct assessment’. The experimental aims in development of the FAPs were to: i define HGMUs at each site; ii make regular measurements of physico-chemical variables; iii study specific processes; and iv examine relationships among wetland properties and characteristics that could be used for predictive purposes. The empirical information collected, supplemented by the available literature and consultation with international interdisciplinary experts, forms the technical basis of the FAPs (Figure 23.3). Clear relationships were found to exist between individual HGMUs and specific wetland functions, including nutrient removal and retention (Baker and Maltby 1995; Russell and Maltby 1995), floodwater control (Hooijer 1996), ecosystem maintenance (Clément et al.

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Table 23.3 Example assessment of an enclosed water body’s suitability for swimming. The numbers in the first 3 columns indicate the questions that were answered ‘yes’. Notation (e.g. #, (2|3)) refers to incomplete information about the site, and is explained fully in the procedures document. CV1 Water Conditions Q1 Can the HGMU be described as a lake, pond, or reservoir? Q2 Is the HGMU more than 1.5 metres deep? Q3 Is the temperature of the water between 15°C and 25°C? CV2 Evidence of Suitability for Swimming Q1 Is the water free of contaminants, such as sewage or agricultural effluents? Q2 Is there an absence of dangerous wildlife (e.g. snakes, leeches)? Q3 Is the substrate on the bottom of the HGMU sand? CV3 Evidence of use for Swimming Q1 Has the HGMU ever been used for swimming? Q2 Have swimmers been free of adverse health effects? Q3 Are there any facilities such as lifeguard post, changing rooms, toilets around the HGMU?

CV1

CV2

CV3

Rationale

Code

123

123

123

1/1

123

123

−

123

1(2|3)

−

123

3

−

#1

−

−

1 #(2|3)

−

−

All questions are answered with ‘Yes’ and it is certain that the HGMU provides suitable conditions for swimming. All questions in CV1 and CV2 are answered with ‘Yes’ but the answers in CV3 are unknown or uncertain. It is certain that the HGMU provides suitable conditions for swimming but there is no evidence of being used for swimming previously. It is therefore recommended that swimming occurs only with some caution. All questions in CV1 are answered with ‘Yes’. The questions in CV2 are answered as: ‘Q1 is true and Q2 or Q3 is also true’. It is therefore certain that the HGMU provides suitable conditions for swimming. However, there may be dangerous wildlife or an unsuitable substrate and so swimming should occur only with considerable caution. All questions in CV1 are answered with ‘Yes’, however, only question Q3 in CV2 is definitely true. It is therefore likely that the HGMU provides suitable conditions for swimming, as the morphology of the area is appropriate for swimming, but it might be dangerous owing to poor water quality or dangerous wildlife. Swimming is recommended only with extreme caution. The answer to CV1, Q1 is negative. The HGMU is therefore not a water body appropriate for swimming. The answers to other questions are irrelevant in this context. The answer to CV1, Q1 is positive but the answers to Q2 and Q3 are unknown or uncertain: the HGMU is a water body appropriate for swimming but is either not deep enough or does not have a suitable temperature. The HGMU is therefore probably not fit for swimming and the answers to the other questions are irrelevant.

1996) and food web support (Castella and Speight 1996). Validation of the HGMU approach was essential given the underlying importance of the functional unit in the prediction of distinct patterns of processes and functions within wetlands. This was undertaken using a variety of statistical methods and approaches, with outcomes reported in McInnes et al. (1998).

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1/2

1/3

1/x

4 2

In the case of the river marginal wetlands, the field investigations identified three important characteristics: • The importance of environmental gradients across the study sites; • Strong separation between slope and floodplain units (owing to basic hydrological differences); and

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Ecological, hydrological and biogeochemical process studies

Identification of key wetland processes

Laboratory experiments

Expert knowledge and literature resources

Field data collection Field experiments Definition of European wetland functions

Socio-economic studies

Analysis and evaluation of data

Impacts of human activity upon functioning

Hydrological and ecological modelling

Characterisation of functions and identification of predictors

Functional Analysis system development

GIS and catchment scale studies Initial user group consultation

Operationalisation Hydrogeomorphic classification

Development of European wetland functional analysis procedures: activities and resources

User legislative and regulatory requirements European wetland functional analysis procedures for : River marginal wetlands (FAEWE I and II, PROTOWET) Lake marginal and estuarine wetlands (PROTOWET)

Scientific field testing User group field testing User time and resource constraints

Fig. 23.3 The activities and resources used in the development of the FAPs.

• Differences between floodplain subunits (level, elevated and depressional), reflecting previously unaccounted differences in flooding regime, proximity to the main channel, and soil processes. Using the FAPs to improve decision-making Application of the FAPs enables predictions to be made of how a wetland, or particular parts of it, are working. This allows decisions likely to adversely affect functioning and provision of ecosystem services to be managed. It may be that,

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in some cases, unavoidable impacts can be zoned to minimise loss of particular functions and degradation of services. Assessment of wetland functioning can vary considerably depending on the target objectives of the user, who can choose to adopt a variety of possible approaches. The FAPs are structured to enable: • Overall assessment of wetland functioning; • Assessment of one or more functions; and • Assessment of one or more physical, chemical or biological processes that underpin functioning.

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Functional Assessment of Wetlands Central to the procedure is the gathering of baseline data about the site to be assessed and the adjoining areas of land that influence the HGMU, particularly through hydrological connections. This database forms a background against which information on processes occurring within the HGMU can be assessed. Functions are themselves delivered through the operation of natural processes, one or more of which, rather than the overall function, may be of specific interest to the user. If the user is interested in making a full assessment of the HGMU, then all sections of database establishment and functional assessment need to be undertaken. Alternatively, the user may wish to assess only a single function or a selection of functions, in which case the FAPs are structured so that only information appropriate for assessing a particular function needs to be collected. For example, the user might ask ‘is there likely to be significant loss of nitrate by denitrification?’, rather than ‘is there overall export of nitrogen from the system?’. In this case, the FAPs enable the identification of the appropriate processes, and indicate to the user the need to complete those sections assigned to the selected processes.

FRO M IN F OR M AT IO N T O S IT E AS S E S S M E N T The FAPs manual (Maltby 2009) addresses 38 hydrological, biogeochemical and ecological processes contributing to 12 functions that may occur in wetlands. These are listed in Table 23.1. Key wetland processes have been identified through the empirical research programme, and investigated individually for their effects on functioning. Their recognition may help to increase awareness of the interactions that can occur, and illustrate the significance of maintaining the appropriate conditions to ensure that a function is maintained. Each wetland function is a composite outcome of certain physical, chemical or ecological characteristics and processes, any of which theoretically can be assessed for the degree to which it operates.

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An unaltered wetland in a semi-natural catchment may perform a wide range of hydrological, biogeochemical and ecological functions, whereas one situated in a degraded environment, polluted with water-borne contaminants and suffering physical loss of habitat, might have lost much of its capacity to support wildlife, its ability to provide clean water downstream, and its role in supporting production of consumable animal and plant products. Decisions on how best to improve the degraded system will depend on identification of the critical processes governing functional differences between the two, and the environmental criteria that support them. Functional assessment can assist the necessary scientific guidance. The FAPs can be carried out manually or by using a CD-based electronic version. In both cases, the work involves preparatory site visits and empirical measurements and estimates. Internal structure The FAPs text takes the user through the separate steps, assuming minimal prior knowledge, in order to assess the aspects of functioning that are of interest. The functional assessment is a threestage process: first, the areas of hydrogeomorphic homogeneity are delineated and characterised, then the operation of the processes that govern functioning are identified and evaluated, and finally the assessment is made. To do this, the user takes the following steps: i identify the area to be assessed for functioning: the assessment area (AA); ii establish the baseline conditions of the site (slope, hydrology, soils); iii define the HGMUs using criteria of homogeneity within and dissimilarities with outside; iv characterise each HGMU for physical and chemical parameters (this establishes a database; the user then selects a required function); v answer process-specific questions based on observable characteristics (the user continues with the next process until all processes relevant to the chosen function are examined);

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vi record process scores on an appraisal sheet; and vii the user moves to the next function and continues as above. When all required processes and functions have been investigated, the scoring stage begins. First scores are obtained for processes, next for functions and third for the entire HGMU. When assessments for all HGMUs are complete, the appraisal sheets are combined to obtain scores for a specific function across all HGMUs within the AA, for all functions across the HGMU, or for all functions across the AA. There are distinct stages in the application of the procedures (Figure 23.4). Stage One The assessment area (AA) is that part of the wetland that is of particular interest. Typically, the AA will not be homogeneous but will be characterised by local changes in slope, hydrology and management activity across the site. In the first stage of the procedure, the AA is examined for similarities and dissimilarities within its boundaries. This can be done from aerial photographs or by physical examination of the site, but includes detailed mapping of the HGMUs within the AA and the area around it that is likely to influence functioning of the AA. In either case, wherever possible, decisions on the characterisation and differentiation of the AA should be verified in the field and, where appropriate, by use of local knowledge and secondary sources of information. This initiates establishment of the database that will form the backbone of the FAPs for the site. The AA characterisation will reveal distinct areas with suites of similar hydrogeomorphic features that are distinct from neighbouring areas. These are the HGMUs discussed above, and are the basic units of assessment. Each will have discrete characteristics that influence functioning in particular ways. A wetland site at Kismeldon Meadows, Devon, UK is a typical river marginal wetland of wet grassland, and provides a good example of the operation of the first stage of the procedures in

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practical terms. The area to be assessed is drawn as a simple base plan for use in the field. An aerial photograph or a photocopy of an existing map can be used as a template. The topographical features of the assessment area and adjacent contributory areas (CA), in this case the river and surrounding fields, are represented on it. A standard colour guide is used for adding details to the base map. Figure 23.5 shows the progressive development of the base map as information is obtained from the desk and field studies. The site investigation can usually be done in a day by walking transects across the AA. Records of the site, conditions and personnel are identified on the sheet, and site and management information important for database establishment are added directly on the map. Local people with knowledge of the site are a very valuable resource since they can often provide information that is not available from maps, field surveys or archives, especially on issues of management and flooding. A profile of the site at Kismeldon Meadows, showing HGMUs on slopes and the river floodplain, can be seen in Figure 23.6. The final map shows relevant physical and management information including, for example, boundaries, rivers, ponds and drains, details of substrate, extent of flooding, management such as ploughing or burning, habitat status and so on. (Figure 23.5d). The baseline information is critical to effective assessment of functioning because processes are modified by the character and management of both the assessment and contributory areas. Once HGMUs are delineated, their defining features of geomorphology, hydrology and ecology are quantified where appropriate and recorded for addition to the database. These include establishment of the patterns of land use and management, estimation of HGMU area, the assignment of codes for hydrology and the determination of inputs to the HGMUs of key nutrients (nitrogen and phosphorus) and trace elements that may be important. The HGMU-specific database is used to estimate the operation of selected physical, chemical and biological processes that will be made

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HGMU

Characteristics

HGMU

Characteristics

HGMU

Characteristics

Database establishment

Parameters of AA

Process Process Process

Interpretation

Interpretation Interpretation

D A T A

Process Process Process

Degree of function operation (summing)

Degree of operation (process) S E

(Choose function)

Test controlling variables (ask process questions)

A

2

Define AA

Characteristics

B

1

HGMU

Definite

1/1

Part

1/2

Unknown

1/x

Not significantly

2

Not

4

Interpretation Interpretation

Interpretation Interpretation

Appraisal sheet

Interpretation Interpretation Interpretation

Cumulative function score

Next function

3

Functions and scores

HGMU 1 1 2 3 4 5

Assessment: one function across AA

n

HGMU 2 1 2 3 4 5

Assessment: functions for entire AA

HGMU 3 n

1 2 3 4 5

n

Assessment: functions for one HGMU

Fig. 23.4 Functional assessment is conducted in three stages, each consisting of several steps that compile information. Stage 1 creates a database against which process-specific variables are evaluated; stage 2 asks processspecific questions so that the operation of processes can be determined; stage 3 compiles the information of previous stages into an assessment of functioning dependent on the requirements of the user.

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Functional Assessment of Wetlands Floodplain Unit

Gentle Slope Unit

Floodplain Unit

Hydrogeomorphic Unit Moderate Slope Improved Grassland

Gentle Slope

Floodplain Humid Grassland

Wet Health S = Soil A = Alluvium and head WZ = Heavily weathered rock FB = Fractured rock

Dry Grassland

1–5 = Instrumented sites = Groundwater flow

400 m

KMSL 1/1 R. Torridge

WZ

FB

KMSL 2/2

S

10 m

KMSL1/3

Piozometric surface

KMSL1/4

KMSL1/5

Seepage Area

Low permeability unit Gleysols (Poorly drained, slowly permeable, silty)

Silty fluvisols Poorly Well drained drained

Fig. 23.6 A profile across the assessment area of Kismeldon Meadows, Devon, UK, showing broad landscape divisions, hydrogeomorphic features and (points 1–5) the location of experimental sites. (Maltby et al. 1996.)

in the second stage of the functional assessment procedures. Stage Two This allows the user freedom to assess any aspect, from the performance of a single process or function, to a full assessment of all functions potentially operating in each HGMU. The assessment of each process requires the user to find answers to predictive questions about indicators, which reveal the behaviour of the controlling variables

(CVs) that themselves influence each wetland process. When combined and qualified by relevant values in the database established in stage one, these answers progressively identify the features governing the operation of processes, and so strengthen predictability of process performance. Notes, provided at each step, discuss significance and provide guidance for interpretation. Figure 23.7 shows, for just one example, the relationships between the database, the CVs, the processes evaluated and the function assessed from the combined outputs.

Fig. 23.5 Development of a site base map from desk and field-based studies: (a) the assessment area (AA) is delineated on a simple map showing the obvious features of the site including aspect, boundaries and major topographical features; (b) details of smaller streams, drains and ponds are shown in blue, and the lines of transects are placed on top of these; (c) details of management, flooding and other relevant information for the AA (and the CA) are overwritten, and (d) (simplified to aid clarity) positions and areas of the HGMUs are marked and numbered. Codes shown (e.g. improved grassland is indicated as IG) are described fully in the FAPs document. (Maltby 2009.) Basic fieldwork and map information are added in a box below the map.

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Assessment for the process of denitrification contributing to the ‘nutrient export’ function Database

Controlling variables

Processes

Function

Controlling variable: nutrient input

Controlling variable: soil carbon

Example for one CV Indicator: nutrient inputs

Indicator: soil profile

Indicator: vegetation type

Controlling variable: soil pH

Controlling variable: potential for interaction with nitrate

Controlling variable: soil oxygen status

Controlling variable: soil temperature

Process: Gaseous export of N: ammonia volatilisation

Process: Gaseous export of N: denitrification Function: nutrient export Process: Gaseous export of N: export through land use

Process: Gaseous export of N: export through physical processes

Fig. 23.7 An example of the decision tree in assessment of the process of denitrification in a wetland HGMU using the FAPs. The process is part of the function of nutrient export. The database has previously established details of nutrient inputs, soil profiles and vegetation types, and the responses to the CV questions are estimates or measurements of current conditions in the HGMU. Together, these values are used to assess the performance of the relevant process. Subsequently, the outputs can be combined with those of other processes operating within the function to give an overall indication of probable function performance, which in this example, may contribute to the ecosystem service of water quality.

Once all the CV questions have been answered for a particular process, the answers are compared with the combinations of answers in look-up tables until a match is found. When all CV answers have been matched to their counterpart in the table, the assessment outcome will have been determined. Depending on the number of controlling variables defining the process, and the combination of their different values, there may be numerous possible performance outcomes, but

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the outputs are grouped according to four basic conclusions: • the process is definitely performed – code 1; • the process is performed but only to a small degree, or there are unknowns – code 1/x; • the process is not significantly performed – code 2; and • the process is definitely not performed – code 4. A rationale statement, describing the assessment, accompanies each output code. Where the output for the process is Code 1 (definitely

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Functional Assessment of Wetlands performed) there are likely to be degrees of expected performance, for example the process may be definitely performed but not to the extent possible had some of the controlling variables been different. Accordingly, the output code can be modified to reflect this, so that a code of 1/1 would be given to the highest performance of the process, and codes of 1/2, 1/3, … 1/n. would be given where the assessment dictates a progressively lower performance of the process. This enables greater flexibility in the assessment so that subtleties of process performance are revealed. In addition, some processes, such as estimation of the retention or export of a specific nutrient, permit a degree of quantification. For these processes, as well as the rationale statement and code, a range of measurable values expected for the process is given, and the suffix code for definite performance of the process is reported as 1/a, 1/b, … and so on. The rationale statement that accompanies each output code describes the interpretation of the process outcome. This series of statements, based on expert interpretations of the answers found, describes the situation in the HGMU as accurately as possible. Slightly different answers to the CV questions will give subtly different statements of process operation. The outcome codes are recorded on paper or stored electronically. Table 23.3 gives an extremely simplified example of the method used to determine the performance of a process from the answers given to a set of CV questions. These are answered according to the best available information, gained from local knowledge, written records and site visits. Answers in the affirmative are noted and the combination of answers for all CVs is located on the accompanying look-up table. The example has been devised as an illustration only and does not feature in the FAPs. Taking the form of an assessment of the suitability of a body of water for swimming by adults, the questions relate to controlling variables (water conditions, evidence of suitability and evidence of having been used) and, in combination, can evaluate the suitability of the site for the activity (process) of swimming. At the end of the section, all the process output codes are recorded on the appropriate sheet, from

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where they are later combined to give a score indicating the status of the overall function (which, in the case of this illustration, might be ‘recreation’, and contributing to ‘cultural’ ecosystem services). In the electronic version of the FAPs, these appraisal sheets are completed automatically as the answers are entered into the computer. Stage Three Stage three of the FAPs uses the summary process information to determine overall scores for each function. When the score for a particular process is established, it cannot be simply added to scores for other processes, because some processes might be more important than others. For example, the importation of biomass into the HGMU by wind transport, which may operate at a high level, may not be as significant as the increase in biomass within the HGMU owing to primary production. Accordingly, the scores for these two processes within the function of food web support may be weighted differently in order to reflect the particular situation in the wetland. The combination of individual process scores to derive a score for overall functional performance is based on the allocation of numerical values between 0 and 1 to each output code. Codes in the range 1/1, 1/2, …, 1/n, and 1/a, 1/b, 1/c are allocated higher values (e.g. 0.8), based on their relative strengths of operation and the extent of output codes for the process, and codes 1/x, 2 and 4 are given low (e.g. 0.25) or zero values. These numerical values are multiplied by factors that are weighted according to their relative contribution of the process to the function as determined by expert judgement, in order to give a score that is a balanced reflection of the importance of the process. When final values have been assigned to all processes comprising a function, a simple mean is taken in order to derive a value for the operation of the function as a whole in that HGMU. If the user has worked through the entire FAPs, values for all functions in the HGMU will have been obtained (Figure 23.4, stage 3), thus a picture of

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overall functioning in the HGMU will have been obtained. These can be combined with the scores of other HGMUs to arrive at a mean score for one function across the AA, or to obtain an assessment of the general state of wetland functional performance in the AA. Where there are differences in area between HGMUs, these might be used to further weight individual HGMU scores, so that a fair representation of functioning can be obtained over the entire area of assessment.

A N O P E R AT IO N AL AID T O W E T LAN D M AN AG E M E N T The FAPs enable the decision-maker to evaluate the functions and services provided by a wetland, and can be used for objective comparisons of functional performance before and after changes in management strategy, but they can also be used predictively. By making changes to the database or CV answers, the user can perform a theoretical trial of potential management options under different scenarios. For example, it may be possible to change hydroperiod by use of a bund or sluice gate. If the information entered in the FAPs is changed to reflect this, an indication of the probable functional outcomes will be obtained. This scenario-testing facility is most easily done on the electronic version. The maximum benefit of the FAPs can be realised when using the electronic version in combination with a downloadable management evaluation system known as the Wetland Evaluation Decision Support System (WEDSS) (Mode et al. 2002; WEDSS 2007), which complements the FAPs, and supports the appraisal of management practices. The WEDSS capacity for scenario building allows the user to ask, ‘What if?’ for a variety of management options. For example, if it was suggested that flooding an HGMU would provide improved habitat for birds, the WEDSS would permit assessment of the effects of this management change on other functions such as nutrient retention and flood detention. In this way it is possible to present numerous scenarios to the WEDSS, using the

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resulting assessments to decide on appropriate action for the wetland and catchment. In addition, the site-specific database and maps, along with the FAPs, are held within a GIS (Geographical Information System) framework that also holds a multi-criteria analysis (MCA) tool. This enables evaluation of wetland management proposals whilst taking account of social and economic criteria that have a bearing on management. Outputs are in the form of annotated maps and associated data, accompanied by detailed guidance for the functional relevance of proposed changes. This gives the user the ability to review the potential impacts of any proposed management changes as biophysical effects on the wetland assessment area. The MCA facility helps the user to choose appropriately from several alternative options. By using a dimensionless scale, MCA analyses the options by direct comparison of different inputs, whilst integrating environmental, social and economic criteria. The MCA output is a single value representing a standardised total score for each alternative. This offers increased flexibility of the package because the alternatives can be HGMUs, groups of HGMUs, the entire assessment area or alternative management scenarios. It thus allows targeting of HGMUs where management action may be required, as well as permitting cumulative comparisons to be made of alternative management scenarios. An assessment was made using WEDSS in the catchment of the River Tamar, UK, to assist evaluation of management options in a riparian wetland with floodplain and slope HGMUs. There were three potential management scenarios for the site according to the following criteria (Table 23.4): Degraded – current management and development policies that adversely affect wetlands continue unchanged. Policy Compliance – where environmental policy targets are achieved. Organisations like the UK’s Environment Agency produce local plans for water resources, which address management options that will lead to greater sustainability. These plans outline targets for such aspects as water quality and supply, and environmental

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Functional Assessment of Wetlands improvements associated with more optimised resource use in the future. Green – this scenario assumes a shift in societal values, and focuses on sustainable lifestyles. Community-level actions are important, and all decisions are informed by a strong emphasis on education and participation. Driving all decisions Table 23.4 Management of the culm (wet grassland) landscape at the Tetcott Barton site under different environmental change scenarios. Scenario

Management

Degraded

Conversion of existing culm-like area (2.3 ha) to intensive agriculture Current situation – existing culm-like area not fertilised and grazed lightly Restore adjacent area to culm-like status (additional 0.35 ha)

Policy compliance Green

is the knowledge that human actions are inextricably linked to the state of the environment, and the aim of restoring the well-being of the environment becomes the highest priority. The initial stages of the FAPs were applied as described above. The answers to the CV questions can be entered are inputted directly into the WEDSS, which automates the FAPs assessment processes for each HGMU to derive a spatiallydistributed assessment of wetland functioning across the AA. Figure 23.8 shows an example of the WEDSS outputs for one process (sediment retention) annotated with the output codes and rationale statements applicable to each HGMU. To represent the three scenarios, the field and desk study data that describe the HGMUs were amended by expert judgement to reflect the probable status of each in the event that the three management scenarios were implemented. The answers to the CV questions were subsequently

4/1: There is evidence that the function is not performed.

1/3: The HGMU retains suspended sediments from river/lake water during frequent floods, but it is unknown if this forms a significant proportion of the suspended sediment load in the river/lake. A lack of erosion-indicators suggests that the retention will be long-term.

1/13: The HGMU retains suspended sediments from river/lake water during frequent floods, but it is unknown if this forms a significant proportion of the suspended sediment load in the river/lake. Some erosion-indicators suggest that the retention may be short-term.

1/7: The HGMU is very likely to retain suspended sediments but it is unknown if the sediment source is river/lake water or surface runoff. A lack of erosion-indicators suggests that the retention will be long-term.

Legend Code 1/3

1/6: The HGMU is likely to retain suspended sediments from river/lake water but flooding is rare or the flooding frequency is unknown. A lack of erosion-indicators suggests that the retention will be long-term.

1/6 1/7 1/13 4/1

Fig. 23.8 The WEDSS output map for the Sediment Retention evaluation at Tetcott Barton.

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adjusted to reflect the probable HGMU status under each scenario. The WEDSS produced three sets of functional assessment maps, representing the expected performance of the selected wetland functions under each management option. The output codes in the electronic FAPs for each process are assigned numerical scores to facilitate analysis using MCA. This allows the functional assessment outputs to be grouped into categories that reflect particular ecosystem services. The WEDSS outputs take the form of maps with active annotation. Placing the cursor over the map displays the rationale statements that explain changes for each ecosystem service considered. The WEDSS combines individual process scores to indicate the performance of ecosystem services for each of the three scenarios. It therefore allows an overall assessment of the AA, the contribution it makes to the range of ecosystem services and how that contribution may change under the different management options. In this way, the WEDSS provides a robust and objective means of supporting decision making. For example, in Figure 23.9, the map sequence that indicates performance of phosphorus cycling, contributing to the ecosystem service water quality, reveals the expected response of phosphorus (P) concentrations to management changes in the three scenarios. The reduced application of P for agriculture under the Green scenario is reflected in the output maps, which show that the retention and export of P decreases with improved ecosystem status. Correspondingly, its higher application under the Degraded scenario is revealed as increased P retention and export. Similarly, faunal diversity is shown to respond positively to wetting and reduced agricultural intensity in the Green scenario when compared with the Degraded and Policy Compliance scenarios. These assessments of ecosystem services provide an overview of the assessment area for management purposes but, if required, a user of the WEDSS could look in more detail at the scores for the individual processes and functions underlying any of the ecosystem services. It may be necessary, for instance, to determine the effect of the

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management scenarios on the different elements of the P cycle (e.g. export through vegetation management, retention through plant uptake, retention of particulate nutrients). This case study demonstrates the capacity of the WEDSS for comparing wetland management scenarios under alternative strategies. The WEDSS is designed to be a flexible tool for addressing a range of wetland management applications, including: • Flood risk/flood retention capacity; • Groundwater recharge/discharge; • Protection of surface water quality; and • Maintenance of valuable biodiversity. It can be extended further to include features such as: • Social values (e.g. recreation, education); • Potential for reestablishment and restoration of wetlands; and • Suitability of conditions for archaeological preservation. To assess the effect of wetland restoration on the protection of surface water quality the user would first apply WEDSS to the wetland in its current state. The controlling variables that would be affected by the particular management measure to be adopted would then be identified. The methods used to implement the restoration (e.g. infilling of drainage ditches, bund establishment) will determine the controlling variables and the degree to which they are affected. The interactions among controlling variables, processes and functions are complex, since individual controlling variables affect a number of processes and functions in a wetland. These interactions make it difficult to predict the effect on functioning of a given change in wetland state without a tool such as WEDSS. Extrapolation across landscapes The same three options for management were used in functional assessment of the Humberhead Levels, Yorkshire, UK, where HGMUs in nine assessment areas were compared with comparable areas of land outside the AAs. Extrapolation of the HGMU outputs to closely similar units in the same catchment is an effective method of

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Fig. 23.9 Some selected WEDSS outputs from the Tetcott Barton study site showing changes expected under three alternative management scenarios. Shades are based on the multi-criteria analysis score for each HGMU, and indicate the degree of functioning, with darker being greater. The scores for phosphorus and nitrogen cycling reveal lower levels of functioning under improved environmental management. This reflects the lower input concentrations of the two nutrients under these management scenarios.

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evaluating functional performance on a landscape scale. This former wetland, 2275 km2 in extent, has been mostly drained for agriculture, with just a few remnant wetland areas still surviving. The assessment of the Humberhead Levels involved an examination of the potential social, economic and environmental advantages of a return to sustainable management of the area consequent to a general raising of the water table. The condition of most of the Levels was considered to be degraded and not restorable under prevailing socio-economic attitudes. Existing wetland remnants were evaluated for their potential as targets for restoration and improvement. Nine pilot sites, considered to be closely representative of the major landscape features, were selected for HGMU delineation and functional assessment. Soil and geology maps identified the extent of different landscape types within the Humberhead Levels area, and GIS was used to

identify areas that were functionally similar to the pilot study sites. The outputs of each functional assessment were scaled up to HGMU-similar areas in 5 × 5 km squares surrounding each pilot study site. A second stage of extrapolation was used to infer functional assessment outcomes for the whole Humberhead Levels area (Hogan and Maltby 2005). When assessment outcomes are extrapolated to the wider landscape, there are inevitably losses, for practical reasons, in the level of detail available at the local HGMU scale. Therefore it is important that, if the patterns of wetland types are complex, or doubts exist about the appropriateness of the selected pilot sites as representative of broader landscape units, verification surveys are carried out to establish the validity of assumptions. The three alternative management scenarios represent a progressive increase in the presence of water in the Humberhead Levels. Table 23.5 gives

Table 23.5 The three alternative land use scenarios for the proposed options for re-wetting of the freshwater parts of the Humberhead Levels, Yorkshire, UK. (From Hogan and Maltby 2005.) Landscape scale Degraded HGMU

Policy compliance

Green

River alluvial flats

Drains and ditches blocked to raise water levels and convert any arable land to permanent pasture.

Embankments removed or set back to enable flood pastures to extend across the floodplain.

Sand lands affected by groundwater

Clay lands (on former lake clay deposits) Valley fen peat

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Permanent pasture on alluvial gley soils, some sections embanked, otherwise flood pastures; occasional spring cereals where drained. Pump-drained, arable farming.

Altered pumping regime to allow Surface water penned in ditches and the water table to rise and provide drains blocked; wet grassland and groundwater irrigation for damp fen established. grassland; some blocked ditches and aquatic ecosystems established. Drained for dry/damp grassland; Drains blocked for wet grassland. Drains blocked and surface water clayey stagnogley soils. penned in ditches; wet grassland and fen with aquatics in ditches. Wetland nature reserves where Wetland nature reserves where land Wetland nature reserves where land land disturbed or subsided – disturbed or subsided – otherwise disturbed or subsided – otherwise otherwise drained for less intensively drained, mainly wet surface water penned in ditches arable; organic soils, peat grassland; organic soils, peat locally and drains blocked, fen or fen-carr locally stabilised but mostly stabilised, otherwise continued slow established; organic soils, peat continuing rapid degradation. degradation of peat. stabilised with some growth locally.

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Functional Assessment of Wetlands an approximation of the management strategies of the three scenarios, and Table 23.6 shows the extrapolated assessment outputs for the whole of the Humberhead Levels according to major landscape types for the three scenarios and the hydrological, biogeochemical and ecological functions assessed by the FAPs. Field testing The FAPs have been subject to comprehensive field testing at a wide range of European wetland sites. A detailed evaluation for three diverse sites can be found in McInnes et al. (1998). The outcome of the trials led to refinements, which included some of the terminology used in the procedures, ways of recognising field indicators of hydrological limits, ways of identifying HGMU components and changes to the ways of obtaining baseline information in the desk survey work. These tests, together with others carried out by scientific partners in the European projects, have highlighted many issues involved in transforming detailed scientific knowledge into practical assessment procedures. In the light of such critical evaluation, the ‘final’ form of the FAPs has been subjected to considerable detailed revision. This process inevitably will continue as user feedback and experience is accumulated. The FAPs should be considered as a work in progress, providing a ‘best estimate’ substitute for empirical research but nevertheless providing an evidence-based support to management decisions.

H O W T HE F AP S CAN CON T R IB UT E TO B E T T E R M AN AG E M E N T O F W E T L AN DS The FAPs aim to provide objective evaluation of wetland functioning by experts or non-experts to support better decision-making at the management level. They make obvious the direct connections between the integrity of wetland processes, the wider impacts of these in relation to ecosystem services valued by human communities and the management and policy decisions

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that influence ecosystem functioning. Integration with the WEDSS allows choices to be made concerning the use of the wetland site directly, and management at wider spatial scales, including whole catchments. Effective catchment management can benefit from a landscape-wide application of the FAPs, which may be achieved by combining the outputs of different HGMUs and extrapolating them over broader landscape areas. The method is to identify areas in the catchment that share similarities with the HGMUs that have undergone functional assessment. The assessment outcomes of these can be applied on greatly broadened scales. This is especially useful when evaluating alternative management options for an area, for example where the option of continuing existing management patterns (‘degraded’) is compared with an option featuring greatly enhanced environmental criteria (‘green’), or management which meets statutory policy guidelines (‘policy compliance’). Use of the FAPs has the simple requirement of a general knowledge of the natural environment, a few basic pieces of equipment, and time available for carrying out the assessment. Users are guided through the procedures step-by-step, and any unfamiliar skills and techniques usually can be learnt along the way. In just a few places, an indication is given of where more specialist advice or expertise may be required. The FAPs may be used by a range of individuals or organisations who are concerned with wetland management and wish to gain a better understanding of the wetlands for which they have responsibility or interests. A key benefit of using the FAPs is that site management strategies can be better aligned, adjusted or modified to meet specific objectives. For example, management of the water table could enhance fluctuations to increase the rate of denitrification in response to suspected elevations in nitrate runoff from adjacent agricultural land. The FAPs also potentially have a role to play in supporting environmental policy. A typical policy objective might be the decision to create wetland buffer zones alongside river channels in order

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NA NA N N P P D D N N D D N N P P N N N N

N N N N

Biogeochemical 3.2.A.1 Long-term retention of nutrients by plant uptake 3.2.A.2 Storage of nutrients in soil organic matter 3.2.A.3 Adsorption of N as ammonium 3.2.A.4 Adsorption and precipitation of P in soil 3.2.A.5 Retention of particulate nutrients (N and P) 3.2.B.1.1 Gaseous export of N by denitrification 3.2.B.1.2 Gaseous export of N by ammonia volatilisation 3.2.B.2 Export of nutrients (N and P) by vegetation management 3.2.B.3 Export of nutrients by wind and water 3.2.C.1 Organic matter accumulation

Flat areas N N N N

Uniform slopes

Process/function Hydrological 3.1.A Flood detention 3.1.B Groundwater recharge 3.1.C Groundwater discharge 3.1.D Sediment retention

Functional units Undifferentiated depressions NA N P D N D N P N N

N N N N

Ditches Lakes N N N N

P N N N

River alluvial flats

NA NA NA N N N N N P D N D P N N P D D N N N N N P N N N N N N

D P P N

Flat areas P N P D N D D P N P

N N N N

Uniform slopes P N P D N D D P N P

N N N N

P N P D N D D P N P

N N N N

Undifferentiated depressions

Policy compliance

P P N N N P P P N P

D N D N

Ditches

Degraded

Narrow floodplains of minor streams P N P D D D D P D P

D N N D

Flat areas P P P D N D D D N P

N N N N

Uniform slopes P N P D N D D D N N

N N N N

Green

P D P D N D D P N D

N P P N

Undifferentiated depressions

Land management scenarios

P P N N N P P P N P

D N D N

Ditches

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P N P N D D D D N N

D N N D

Narrow floodplains of minor streams

Table 23.6 A compilation of the Humberhead Levels FAPs outputs for groundwater impacts on areas of sand lands based on the three scenarios of Degraded, Policy Compliance, and Green. D = The function is definitely being performed; P = The function is probably being performed but there are constraining factors or uncertainties; N = The function is not, or not significantly, being performed; NA = not available. Where a function is definitely or probably being performed, it is also shaded in the table of outcomes, so any patterns in functional performance can be more easily seen in the tables of outcomes. (Adapted from Hogan and Maltby 2005.)

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Ecological 3.3.A.1 Provision of overall habitat structural diversity 3.3.A.2.1 Microsites for macroinvertebrates 3.3.A.2.2 Microsites for fish 3.3.A.2.3 Microsites for herptiles 3.3.A.2.4 Microsites for birds 3.3.A.2.5 Microsites for mammals 3.3.A.3 Provision of plant and habitat diversity 3.3.B.1 Productivity/biomass production 3.3.B.2.1 Biomass import via watercourses 3.3.B.2.2 Biomass import via overland flow 3.3.B.2.3 Biomass import via wind transport 3.3.B.3 Biomass import via biological processes 3.3.B.4.1 Biomass export via watercourses 3.3.B.4.2 Biomass export via overland flow 3.3.B.4.3 Biomass export via wind transport 3.3.B.5.1 Biomass export via fauna 3.3.B.5.2 Biomass export via anthropogenic means N N N N N P N N N N P P

N N N N N P N N N N P P N N P N N

N N N N N P N N N N P P N N P N N

P P D P N P N D P N P P D P P P N

P N D P N P N D N N P P D N P P N

N N N N N P N N N N P P P P P N N

P N N N P D D D N N P P P D P N D

P N N N P D D D N N P P P D P N D

P N N N P D D D N N P P P N P N D

D D P P P D D D N N P P N N P P P

P N N N P D D D D N P P D D P N D

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D

N P P P P D D D P P P P P N P

N P P P P D D D P P P P P N P N D

N P P P P D D D P P P P P N P N D

D D P P P D D D N N P P N N P P P

P D P D D D D D P D P D P P P D D

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to enhance biodiversity and to reduce diffuse pollution from agricultural fields. It may be assumed that the part of the floodplain directly adjacent to the main channel is the obvious site for wetland creation, but very often this area does not have the characteristics most suited to pollution control. For example, functional assessment of a site in Devon, UK, identified the most effective sites for installation of buffer zones to be sloping areas at right angles to the river channel, and not the first intuitive position adjacent to the riverbank (Blackwell et al. 1999; Blackwell et al., Chapter 19). Through assessment of overall wetland functions and their contribution to factors such as water quality, water quantity and biodiversity, the FAPs can assist strategic planners at local and catchment scales to make decisions on wetland management and land use. Knowledge of functioning can be used to help with the implementation of national and supra-national policy, such as the Water Framework Directive. In this context, wetlands may be used to support basic or supplementary measures to enhance or maintain water quality and other aspects of ecological condition.

CON CL US ION S The FAPs are the culmination of many years of research and refinement. They provide a simple way of gaining objective assessment of wetland functions and their degree of operation, and provide insight into ways in which management can be put to effective use in rectifying or optimising particular features of functioning. This facility enables assessment of how wetlands work to the benefit of humans and the environment, thus allowing the inclusion of objective determinations of ecosystem service delivery in economic decision-making. This feature represents a way forward in spatial planning, where the roles of wetlands from site to catchment can be more effectively integrated to meet strategic policy objectives. There is still a need for ongoing testing of the procedures, including their

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extension to other functions and improvements in user-friendliness. It is also necessary to link the biophysical assessment of wetlands more effectively with economic as well as socio-cultural valuation. Such development will help to direct the assessment of functioning and derived ecosystem services into the more sustainable management of our wetland resources for the wider benefit of society.

R EFER EN CES Adamus P.R. 1983. A Method for Wetland Functional Assessment, Volumes I and II. FHWA Assessment Method Rep. No. FHWA-IP-82-23/4. Federal Highway Administration, US Department of Transportation, Washington, DC. Adamus P.R., Stockwell L.T., Clairain E.J., Smith R.D. and Young R.E. 1987. Wetland Evaluation Technique (WET), Vol. II. Operational Draft TRY-87. US Army Corps of Engineers, Waterways Experiment Station, Vicksburg, MS. Adamus P.R., Stockwell L.T., Clairain E.J. Jr., Morrow M.E., Rozas L.P. and Smith R.D. 1991. Wetland Evaluation Technique (WET), Vol. I: Literature Review and Evaluation Rationale. WRPDE-2. US Army Corps of Engineers Waterways Experiment Station. Vicksburg, MS, 299 pp. Baker C.J. and Maltby E. 1995. Nitrate removal by river marginal wetlands: factors affecting the provision of a suitable denitrification environment. In: Hughes J.M.R. and Heathwaite A.L. (editors), Hydrology and Hydrochemistry of British Wetlands. Wiley, Chichester, pp. 291–313. Bartoldus C.C. 1999. A Comprehensive Review of Wetland Assessment Procedures. Environmental Concern Inc., St Michael’s, MD, 196 pp. Blackwell M.S.A., Hogan D.V. and Maltby E. 1999. The use of conventionally and alternatively located buffer zones for the removal of nitrate from diffuse agricultural run-off. Water Science Technology 39(12), 157–164. Brinson M.M. 1991. Landscape properties of pocosins and associated wetlands. Wetlands 11, 441–465. Brinson M.M. 1996. Assessing wetland functions using HGM. National Wetlands Newsletter 18, 10–16. Brinson M.M., Kruczynski W., Lee L.C., Nutter W.L., Smith R.D. and Whigham D.F. 1994. Developing an approach for assessing the functions of wetlands.

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Functional Assessment of Wetlands In: Mitsch W.J. (editor), Global Wetlands: Old World and New. Elsevier, Amsterdam, pp. 615–624. Bullock A. and Acreman M.C. 2003. The role of wetlands in the hydrological cycle. Hydrology and Earth System Sciences 7, 358–389. Castella E. and Speight M.C.D. 1996. Knowledge representation using fuzzy coded variables: an example based on the use of Syrphidae (Insecta, Diptera) in the assessment of riverine wetlands. Ecological Modelling 85, 13–25. Clément B., Maltby E., Hogan D.V. and McInnes R.J. 1996. Relationships between vegetation, hydrology and soil properties in river marginal wetlands of the river Torridge basin. In: Merot P. and Jigorel A. (editors), Hydrologie dans les Pays Celtiques. INRA Editions, No. 79, Paris, pp. 305–314. Costanza R., d’Arge R., de Groot R., Farber S., Grasso M., Hannon B., Limburg K., Naeem S., O’Neill R.V., Paruelo J.U., et al. 1997. The value of the world’s ecosystem services and natural capital. Nature 387, 253–260. Defra. 2007a. Securing a Healthy Natural Environment: An Action Plan for Embedding an Ecosystems Approach. Product code PB12853. Defra, London. Defra. 2007b. An Introductory Guide to Valuing Ecosystem Services. Product code PB12852. Defra, London. See also web site http://www.defra.gov.uk/ wildlife-countryside/natres/dea.htm Ehrlich P.R. and Mooney H.A. 1983. Extinction, substitution and ecosystem services. Bioscience 33(4), 248–254. EU. 2003. The Role of Wetlands in the Water Framework Directive. Guidance document No. 12. Common Implementation Strategy for the Water Framework Directive (2000/60/EC). Hogan D.V. and Maltby E. 2005. The Social, Economic and Environmental Benefits of Wetlands in the Humberhead Levels. Stage 3 Report, Part 2: Results and Interpretations of Functional Assessment. Report to the Countryside Agency. Royal Holloway Institute for Environmental Research, Royal Holloway University of London. Hollis E., Holland M., Maltby E. and Larson J. 1988. The wise use of wetlands. Nature and Resources 24(1), 2–13. Hooijer A. 1996. Floodplain Hydrology; an Ecological Oriented Study of the Shannon Callows, Ireland. Dissertation. Free University Amsterdam, Amsterdam. Larson J.S. and Mazzarese D.B. 1994. Rapid assessment of wetlands: history and application to management.

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In: Mitsch W.J. (editor), Global Wetlands: Old World and New, pp. 625–636. Elsevier, Amsterdam. Llamas M.R. 1992. Deterioration of wetlands caused by groundwater exploitation: two case histories in Spain. In: Maltby E., Dugan P.J. and Lefeuvre J.C. (editors), Conservation and Development: The Sustainable use of Wetland Resources. Proceedings of the Third International Wetlands Conference, Rennes, France. September 1988. IUCN, Gland. Llamas M.R., Garcia M. and de la Hera A. 1996. Landscape changes and ecological impacts caused by groundwater abstraction in the Upper Guadiana Basin (Spain). Actas of Proceedings II Paesaggio Culturale Nelle Strategie Europee. Torino, 16–17 May 1996. Lonard R.I. and Clairain E.J. 1985. Identification of methodologies and the assessment of wetland functions and values. In: Kusler J.A. and Riexinger P. (editors) Proceedings of the National Wetland Assessment Symposium. Portland, ME, pp. 66–72. MA. 2005a. Ecosystems and Human Well-being: Wetlands and Water. Millennium Ecosystem Assessment. Island Press, Washington, DC. MA. 2005b. Ecosystems and Human Well-being: Synthesis. Millennium Ecosystem Assessment. Island Press, Washington, DC. Maltby E. (editor) 2009. Functional Assessment of Wetlands: Towards Evaluation of Ecosystem Services. Woodhead Publishing, Cambridge. Maltby E., Hogan D.V., Immirzi C.P., Tellam J.H. and van der Peijl M.J. 1994. Building a new approach to the investigation and assessment of wetland ecosystem functioning. In: Mitsch W.J. (editor), Global Wetlands: Old World and New. Elsevier, Amsterdam, pp. 637–658. Maltby E., Hogan D.V. and McInnes R.J. 1996. Functional Analysis of European Wetland Ecosystems. Phase I (FAEWE). Ecosystems Research Report No. 18, European Commission Directorate General Science, Research and Development, Brussels, 448 pp. Marcotte B.M. 2006. Environmental assessment, CELCO–ARAUCO, and Chile’s wetland sanctuary: ethical considerations. Ethics in Science and Environmental Politics June, 1–4. McInnes R.J., Maltby E., Neuber M.S. and Rostron C.P. 1998. Functional analysis of wetlands: transforming expert knowledge into a practical management tool. In: McComb A.J. and Davis J.A. (editors), Wetlands for the Future, pp. 407–432. Contributions from INTECOL’s V International Wetlands Conference, Gleneagles Publishing, Adelaide.

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Mitsch W.J. (editor) 1994. Global Wetlands: Old World and New. Elsevier, Amsterdam. Mode M., Maltby E. and Tainton V. 2002. WEDSS: integrating wetlands into river basin management to support the implementation of the WFD. In: Ledoux L. and Burgess D. (editors), Proceedings of Science for Water Policy; the Implications of the Water Framework Directive. University of East Anglia, Norwich. Mulsow S. and Grandjean M. 2006. Incompatibility of sulphate compounds and soluble bicarbonate salts in the Rio Cruces waters: an answer to the disappearance of Egeria densa and black-necked swans in a RAMSAR sanctuary. Ethics in Science and Environmental Politics 2006, 5–11. Russell M. and Maltby E. 1995. The role of hydrologic regime on phosphorous dynamics in a seasonally waterlogged soil. In: Hughes J.M.R. and Heathwaite A.L. (editors), Hydrology and hydrochemistry of British Wetlands. Wiley, Chichester, pp. 245–260.

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Semeniuk C.A. and Semeniuk V. 1995. A geomorphic approach to global classification for inland wetlands. Plant Ecology 118(1–2), 103–124. Smith R.D., Ammann A., Bartoldus C. and Brinson M.M. 1995. An Approach for Assessing Wetland Functions Using Hydrogeomorphic Classification, Reference Wetlands, and Functional Indices. US Army Corps of Engineers Waterways Experiment Station, Technical Report TR WRP-DE-10. Vicksburg, MS. Stigler S. 2007. European Commission fights for rare Polish wetland Mega-expressway may threaten endangered birds. Nature online: 9 March 2007; doi:10.1038/news070305–16. UNEP-WCMC. 2007. Doñana National Park, Spain. In Protected Areas and World Heritage, Draft Revision Downloaded 03/10/2007: http://www.unep-wcmc. org/sites/wh/donana.html WEDSS. 2007. Free download: http://www.liv.ac.uk/ swimmer/Past%20Projects.htm

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24 Wetlands Assessment in Practice: Development and Application in the United States Regulatory Context R . DANI EL SMI TH Engineering Research and Development Center, Wetlands and Coastal Ecology Branch, Vicksburg, USA

IN T R O D U CT ION Wetlands are recognised throughout the world as ecosystems worthy of special attention and conservation (Moser et al. 1993; Ramsar 1997). This is due in part to their unique landscape position, the important and often irreplaceable functions they perform and their disproportionate loss historically (Maltby 1986; Conservation Foundation 1988; Dahl et al. 1991). In response to this recognition, a variety of strategies for conserving, managing and restoring wetlands have developed throughout the world. In the United States, a federal regulatory programme has evolved over the last 30 years with the objective of avoiding and minimising impacts on wetlands, through regulating the discharge of dredge and fill-material in ‘waters of the United States’. Many additional programmes also exist at the state and local level to conserve, manage and regulate impacts on wetlands, however the focus of attention here will be the federal regulatory programme, commonly referred to as the ‘404 program’. An effective programme for regulating impacts on wetlands requires a considerable amount of preparatory work. This begins with defining,

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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classifying and making inventories of wetlands, and is followed by developing an understanding of the structural attributes and the processes that characterise different wetland types and their relationship to the larger landscape. This understanding serves as a basis for developing methods to assess the condition of wetlands, manage wetlands for sustainability and restore degraded wetlands. This chapter deals with the latter part of this preparation, and focuses specifically on the development of wetland assessment methods in the context of the 404 program. It provides a brief synopsis of the programme, discusses the challenges of assessing wetland functions in the context of the programme, and provides several case studies that exemplify how current assessment methods are being used to conduct baseline assessments, assess alternatives, evaluate mitigation plans and assess cumulative impacts.

SY N OPSIS OF T HE R EGULAT OR Y PR OGR AMME It is beyond the scope of this chapter to chronicle the sequence of events from which the current 404 program has evolved or to describe the details of programme administration (but see Environmental Law Institute (1993), Cylinder et al. (1995), National Research Council (1995),

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and web link at end). However, it is important to understand how the development and administration of the 404 program has shaped the concept of wetland assessment and its current status in the United States. Consequently, a brief synopsis of the 404 program is necessary. The 404 program is authorised under its namesake, Section 404 of the Clean Water Act (33 United States Code 1344). Text of the Clean Water Act (CWA) is available online (see end). The CWA (33 USC 1344) directs the Secretary of the Army, acting through the Chief of the United States Army Corps of Engineers (USACE) to issue permits for the discharge of dredged or fill-material in ‘waters of the United States’, after notice and opportunity for public hearing. The term, ‘waters of the United States’ (WoUS) generally refers to streams, rivers and lakes, however, wetlands and other special aquatic sites are, by definition, included under the umbrella of WoUS, and subject to jurisdiction under the 404 program (see web link). The 404 program is administered by the USACE in cooperation with other federal agencies that include the US Environmental Protection Agency (USEPA), US Fish and Wildlife Service (USFWS) and Natural Resource Conservation Agency (NRCS). Regulations governing the administration of the programme are outlined in the USACE Regulatory Program Regulations (33 CFR Sections 320–332), and the USEPA 404(b)(1) Guidelines (40 CFR Section 230). These regulations and guidelines are subject to interpretation through Regulatory Guidance Letters, interagency Memoranda of Agreement, and the courts (see web links below). The essence of the 404 program concerns the identification (i.e. ‘delineation’) of wetlands, assessment of impacts on wetlands, and mitigation for unavoidable impacts. Delineation refers to the procedures that are used to identify the extent and exact boundaries of ‘jurisdictional wetlands’ (i.e. those areas subject to regulation under the 404 program) within a proposed project area. Assessment refers to the procedures that are used to evaluate the potential impacts of a proposed project on the functions performed

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by jurisdictional wetlands within a proposed project area. Mitigation refers to avoidance, minimisation and compensation of impacts on wetlands (Council on Environmental Quality 40 CFR 1508.20; Office of Technology Assessment 1984). Wetlands are defined in the 404 program as: ‘… those areas that are inundated or saturated by surface or groundwater at a frequency and duration sufficient to support, and that under normal circumstances do support, a prevalence of vegetation typically adapted for life in saturated soil conditions. Wetlands generally include swamps, marshes, bogs, and similar areas’ (33 CFR 328.3 (b)). This definition provides an adequate general description and characterisation of wetlands, but is too imprecise to establish the extent and exact boundaries of jurisdictional wetlands in a proposed project area. Wetland delineation is conducted through a more rigorous and formal process in which the ‘parameters’ (i.e. criteria) of hydrology, vegetation and soils are sampled at multiple points in the vicinity of potentially jurisdictional wetlands. In order to be considered a jurisdictional wetland, an area must exhibit wetland hydrology, hydrophytic vegetation and hydric soils. The specific procedures currently used to delineate wetlands are outlined in the Corps of Engineers Delineation Manual (Environmental Laboratory 1987). The Natural Research Council (1995) summarises the historical context of delineation and discusses a variety of issues concerning the wetland delineation procedure. The delineation procedure is dynamic and has recently undergone modifications as a result of Supreme Court decisions (i.e., Solid Waste Agency of Northern Cook County v. U.S. Army Corps of Engineers, 531 U.S. 159 (2001) (SWANCC) and the consolidated cases Rapanos v. United States and Carabell v. United States, 126 S. Ct. 2208 (2006)). When jurisdictional wetlands are known or suspected to occur in a proposed project area, project proponents must apply to the USACE for a permit to discharge dredge and fill material into WoUS. Section 320.4 of the USACE regulations summarises the objectives and requirements

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Wetlands Assessment in Practice used by the USACE to determine whether a permit to discharge dredged or fill-material in waters of the United States should be issued. The following excerpt from this section identifies the factors that are considered during permit review: ‘The decision whether to issue a permit will be based on an evaluation of the probable impacts, including cumulative impacts, of the proposed activity and its intended use on the public interest. Evaluation of the probable impacts which the proposed activity may have on the public interest requires a careful weighing of all those factors which become relevant in each particular case. The benefits which reasonably may be expected to accrue from the proposal must be balanced against its reasonably foreseeable detriments. The decision whether to authorise a proposal, and if so the conditions under which it will be allowed to occur, are therefore determined by the outcome of this general balancing process. That decision should reflect the national concern for both protection and utilisation of important resources. All factors which may be relevant to the proposal must be considered including the cumulative effects thereof: among those are conservation, economics, aesthetics, general environmental concerns, wetlands, historic properties, fish and wildlife values, flood hazards, floodplain values, land use, navigation, shore erosion and accretion, recreation, water supply and conservation, water quality, energy needs, safety, food and fiber production, mineral needs, considerations of property owners, and in general, the needs and welfare of the people.’ The specific sequence of steps that is used to review permit applications are prescribed in the 404(b)(1) Guidelines (40 CFR Part 230). They include 40 CFR Part 230, and: 1 Determine whether the proposed project is water dependent, and therefore must be placed in or near a special aquatic site or WoUS; 2 Determine whether practicable alternatives exist for the proposed project; 3 Identify the potential impacts of the proposed project on wetland or aquatic ecosystem

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functions in terms of project specific impacts and cumulative effects; 4 Identify how potential project impacts can be avoided or minimised in terms of project specific and cumulative effects; 5 Determine appropriate compensatory mitigation for unavoidable project impacts; 6 Issue or deny a permit to discharge dredged or fill-material based on an analysis of the benefits gained versus the benefits lost from the proposed project; 7 If a permit is granted, monitor compensatory mitigation to determine compliance. Wetland functions are singled out for particular attention during permit review. Section 320.4 (4) states: ‘No permit will be granted which involves the alteration of wetlands identified as important by paragraph (b)(2) …’ which states: ‘Wetlands considered to perform functions important to the public interest include: 1 Wetlands which serve significant natural biological functions, including food chain production, general habitat and nesting, spawning, rearing and resting sites for aquatic or land species; 2 Wetlands set aside for study of the aquatic environment or as sanctuaries or refuges; 3 Wetlands the destruction or alteration of which would affect detrimentally natural drainage characteristics, sedimentation patterns, salinity distribution, flushing characteristics, current patterns, or other environmental characteristics; 4 Wetlands which are significant in shielding other areas from wave action, erosion, or storm damage. Such wetlands are often associated with barrier beaches, islands, reefs and bars; 5 Wetlands which serve as valuable storage areas for storm and flood waters; 6 Wetlands which are ground water discharge areas that maintain minimum base flows important to aquatic resources and those which are prime natural recharge areas; 7 Wetlands which serve significant water purification functions; and 8 Wetlands which are unique in nature or scarce in quantity to the region or local area.’

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The special significance ascribed to wetland functions in the 404 program Regulations and the 404(b)(1) Guidelines has resulted in the assessment of wetland functions being one of the primary focuses of attention during permit review. There are number of places in the permit review sequence given above where the assessment of wetland functions comes into play. For example, Step 2 requires that impacts on wetland functions be assessed for each alternative, and then compared to determine which alternative will have the least impact. Step 3 requires that wetland functions be assessed and compared under pre- and post-project conditions to determine what project specific or cumulative impacts may result. Step 4 requires that impacts on wetland functions be assessed in order to determine how impacts can be avoided or minimised, and Step 5 requires that appropriate compensatory mitigation for unavoidable impacts be identified. Step 7 requires that wetland functions be assessed and compared before and after mitigation is completed, to determine whether mitigation objectives have been met (Streever 1999). The case studies at the end of this chapter illustrate each of these situations. Another important factor relating to the assessment of wetland functions in the context of the 404 program is the ‘no net loss’ of wetland functions policy. This policy grew out of the National Wetlands Policy Forum Report (Conservation Foundation 1988) which concluded that America’s remaining wetlands were ‘... immensely important to both the environmental and economic health of the nation’, and recommended that federal government policy assure ‘no overall net-loss of the nation’s remaining wetlands’ in the short-term and ‘a significant restoration of wetlands’ in the long-term. In 1990, Congress instructed the USACE to pursue a goal of ‘no overall net loss’ of the nation’s remaining wetlands in Section 307 of the Water Resources Development Act. This policy was adopted by the first Bush administration and became official federal policy in a memorandum of understanding between the USACE and the USEPA (6 February 1990).

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T HE CHALLEN GE OF ASSESSIN G WET LAN DS FUN CT ION S UN DER T HE 404 PR O GR AM There are a number of challenges inherent in assessing wetland functions under the 404 program. These include compliance with the administrative and technical requirements of the 404 program Regulations and Guidelines, dealing with the great diversity of wetlands that exist within the geographic extent of USACE jurisdiction, assessing the spatial discrepancy that can occur between jurisdictional wetlands and ‘functional wetland ecosystems’ (e.g. the spatial extent of jurisdictional wetlands in riparian areas is often considerably less than what is recognised as the functional riparian ecosystem), and deciding how wetland functions should actually be measured, and how to determine whether or not a nonet-loss of wetland functions is being achieved. Each of these issues is addressed in greater detail below. 404 program administrative and technical requirements for assessment procedures The challenge of assessing wetland functions under the 404 program stems in part from the administrative and technical guidelines. These require assessment procedures to be: • Standardised and documented; • Applicable to the variety of wetlands in the geographic extent of USACE jurisdiction; • Applicable in all phases of permit review and follow up (Steps 1–7 above); • Practical to apply given the available time and resources; • Sensitive to significant change in wetland function; • Capable of providing an interpretable and comparable measure of wetland; • Capable of incorporating new technical information as it becomes available; • Capable of adapting to changing programme requirements; • Adaptable for use in other regulatory, management, and planning assessment situations.

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Wetlands Assessment in Practice Taken together, these requirements describe a standardised, documented and flexible assessment procedure that is applicable to a variety of wetland types, practical to apply given available time and resources, and capable of detecting when project impacts or mitigation represent a significant change in wetland function. The administrative and technical guidelines, in and of themselves, do not constitute an insurmountable challenge. In fact, it can be argued that these same guidelines are nearly universally applicable to the majority of assessment situations throughout the world. However, when the time and resources necessary to conduct assessments are considered, the difficulty of the challenge increases considerably. Approximately 1100 USACE employees work in the 404 program. This includes permit processors, enforcement personnel and supervisory and support staff. On average 90 000 permits are processed annually (John Studt, 404 Regulatory Program Chief, personal communication). Several categories of permit exist, based on the extent of the proposed project and the nature of the proposed project. These include ‘letters of permission’ issued for routine types of activities, such as infrastructure maintenance, for which impacts are considered negligible. Nationwide and regional ‘general’ permits, the second category, are issued for activities that are common nationally or in a specific region of the country. They typically involve relatively small areas for which the individual and cumulative impacts have been deemed insignificant. The third permit category is the standard or ‘individual’ permit. This category typically involves larger areas with a potential for greater impacts. The processing of individual permits requires greater time and resources for analysing alternatives, assessing potential impact to wetland functions, identifying ways to avoid impacts, development of mitigation strategies and determining the success of mitigation efforts. The high rate of permit applications means there are not enough staff, time or resources for it to be feasible to conduct intensive quantitative data collection and analysis of wetland functions for all permit applications. At the same time,

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however, it is unacceptable to assess wetland functions using subjective methods that are undocumented or unrepeatable, methods that are insensitive to significant changes in function or methods that do not provide easily interpretable and comparable measures of function. The USACE is well aware of this dilemma, and addressed the issue in the early 1990s in an undated memorandum to the field (see web links). The memorandum recognises that time and resources are constrained, and that the level of analysis required for full compliance with the USACE Regulations and the 404(b)(1) Guidelines may necessarily vary, given the nature and complexity of each individual case. In the context of the different categories of permits, this has translated into an advance determination of no impact or minimal impact, in the case of letter permits and general or nationwide permits. In the case of nationwide permits this determination has been the subject of some controversy (see comments on the proposals for nationwide permits on the web link). This means the majority of the assessment efforts are focused on the standard or individual permit application where impacts are greater and of a more complex nature. To summarise, time and resource constraints resulting from a high permit to staff ratio dictate that the USACE pursue a flexible approach to the assessment of functions, depending on the extent and degree of project impacts. This has resulted in an advance, blanket determination of no impact, or minimal impact, for letter permits and general permits, and focused the majority of the assessment effort on the standard or individual permit category, where the potential project impacts are greater and of a more complex nature. Functional diversity of wetlands Another significant challenge to assessing wetland functions in the context of the 404 program is the wide variety of wetlands that occur across the United States (Mitsch and Gosselink 2000). Wetlands occur as a result of cool or wet climatic conditions, various hydrologic regimes and water sources, impermeable soil conditions and a host

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of other factors. The wide variety of conditions under which wetlands exist leads to the great functional diversity they exhibit. In assessing wetland functions, great functional diversity is problematic because as functional diversity increases, so must the complexity of the assessment procedure and the time and resources required for completing an assessment. There are at least two ways to address the challenge presented by the great functional diversity of wetlands. The first is the ‘one-size-fits-all’ approach, which generalises wetland functions to the point where a relatively small number of criteria can be used to assess the more generic functions performed by all types of wetlands (e.g. provide habitat for wildlife). From a development point of view, the positive aspects of this approach are that fewer functions and assessment criteria need to be identified. From a programmatic point of view, this approach seems to provide a better match with the requirement that assessment procedures be easy to apply given constraints of time and resources. Negative aspects include the fact that assessing a few generic functions usually results in a failure to address some of the more important specific functions performed by certain types of wetland. Another is that often the assessment criteria for generalised functions have low resolution or sensitivity, and often cannot detect significant changes in wetland function resulting from project impacts (Dougherty 1989). Some of the more commonly applied assessment procedures in the United States (Amman et al. 1986; Adamus et al. 1987) use the one-sizefits-all approach. A recent review of more than 40 assessment procedures describes several onesize-fits-all procedures developed over the last 30 years (Bartoldus 1999). Earlier reviews by Lonard et al. (1981); USEPA (1984); and World Wildlife Fund (1992) also provide valuable insights into assessment procedures using this approach. The second way to address the challenge of great functional diversity is a ‘classification approach’. This method begins by grouping wetlands into functionally similar classes, and then identifying the functions and assessment criteria that are

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specific to each class. From a programmatic point of view, this better matches the requirement that an assessment procedure be able to detect significant change in function that results from project impacts. From a development point of view, a negative aspect is that a separate set of functions and assessment criteria must be identified for each class of wetlands. A variety of classifications have been developed for wetlands in North America (Stewart and Kantrud 1971; Golet and Larsen 1974; Millar 1976; Cowardin et al. 1979), and other parts of world (Morant 1983; Semeniuk and Semeniuk 1995). The most widely used system for classifying wetlands in the US is known as the ‘Cowardin’ classification system (Cowardin et al. 1979). This system was developed by the US Fish and Wildlife Service to serve as the basis for classifying wetlands in the National Wetland Inventory (NWI) (Wilen and Frayer 1990). The report that outlines this classification system is available online. While the classification factors used in this system are well suited for wetland inventory using remotely sensed data, they are inappropriate for classifying wetlands based on how they function. For example, forested wetlands in the palustrine, forested, temporarily flooded (PFO1a) category under the Cowardin classification occur in a variety of geomorphic settings and under a variety of hydrologic regimes. However, it should be noted that classification descriptors of landscape position and landform have recently been developed for the Cowardin classification, and promise to make this system more useful in functional assessments (Tiner 1997). A description of this work is online. Brinson (1993; also see Brinson, Chapter 22) developed a hydrogeomorphic (HGM) classification of wetlands specifically designed for use in functional assessment. This classification uses the hydrogeomorphic factors of water source, hydrodynamics, geomorphic setting and other factors of regional importance in grouping of wetlands to identify functionally similar groups, or regional subclasses, of wetland. This reduces functional diversity and simplifies the assessment process by focusing attention on the

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Wetlands Assessment in Practice Table 24.1 Key to the regional wetland subclasses in the Yazoo River Basin, Mississippi, USA. Key characteristics (see discussions on criteria for individual subclasses)

Subclass

1. Wetland is a topographic depression with permanent water or seasonal ponding 2. Wetland is a topographic depression with permanent water at least 2 m deep 3. Wetland is within the 5-year floodplain of a stream 3. Wetland is not within the 5-year floodplain of a stream 2. Wetland is a topographic depression with permanent water 8 ha), including shores subject to seasonal or irregular inundation. • Permanent freshwater ponds (8 ha), including floodplain lakes. Palustrine Emergent • Permanent freshwater marshes and swamps on inorganic soils, with emergent vegetation whose bases lie below the water table for at least most of the growing season. • Permanent peat-forming freshwater swamps, including tropical upland valley swamps dominated by Papyrus or Typha. • Seasonal freshwater marshes on inorganic soil, including sloughs, potholes, seasonally flooded meadows, sedge marshes, and dambos. Forested • Peatlands, including acidophilous, ombrogenous, or soligenous mires covered by moss, herbs, or dwarf shrub vegetation, and fens of all types. • Alpine and polar wetlands, including seasonally flooded meadows moistened by temporary waters from snow melt. • Freshwater springs and cases with surrounding vegetation. • Volcanic fumaroles continually moistened by emerging and condensing water vapour. • Shrub swamps, including shrub-dominated freshwater marsh, shrub carr and thickets, on inorganic soils. • Freshwater swamp forest, including seasonally flooded forest, wooded swamps on inorganic soils. • Forested peatlands, including peat swamp forest. Man-made wetlands Aquaculture/Mariculture • Aquaculture ponds, including fish ponds and shrimp ponds. • Ponds, including farm ponds, stock ponds, small tanks. Agriculture • Irrigated land and irrigation channels, including rice fields, canals, and ditches. • Seasonally flooded arable land. Salt exploitation • Salt pans and salines. Urban/industrial • Excavations, including gravel pits, borrow pits, and mining pools. Water-storage areas • Wastewater treatment areas, including sewage farms, settling ponds, and oxidation basins. • Reservoirs holding water for irrigation and/or human consumption with a pattern of gradual seasonal, draw down of water level. • Hydro-dams with regular fluctuations in water level on a weekly or monthly basis.

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indicators of those benefits. For example, fertility and nutrient characteristics would be crucial in providing forestry and agriculture benefits, but in themselves do not represent benefits (in the anthropocentric sense). Some work has been carried out on converting lists of commonly used terminology and classification of functions to a standardised terminology (Maltby et al. 1995).

SO C I A L W E L F AR E AN D W E T L AN D AS S E S S M E N T It is now apparent that wetlands perform a wide array of functions that can be of considerable value to society. The physical assessment of the functions performed by a wetland is an essential prerequisite to any evaluation of a wetland’s worth to society, but simply identifying these functions is insufficient. Where a wetland is under pressure from human activity that provides measurable economic benefits to society, it will be necessary to illustrate the economic value of the functions performed by the wetland. The provision of such economic information is essential if an efficient level of wetland resource conservation, restoration or re-creation is to be determined. Maintaining a wetland will rarely be entirely cost-less. There will be costs associated with forgoing other uses of the land, or with limiting activities that might impinge upon the ability of the wetland to continue functioning; hence the importance of making explicit the value of the multiple functions that wetlands perform, and of assessing this value within a framework that allows comparison with the gains to be made from activities that might threaten wetlands. This should serve not only to better protect these threatened ecosystems but also to improve decision making for the benefit of society. Economic valuation is therefore a logical extension to any assessment of the functions performed by wetlands for the purposes of public decision making. Comprehensive assessment of wetlands requires the analyst to undertake the following steps: 1 to determine the causes of wetland degradation or loss, in order to improve understanding

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of socio-economic impacts on wetland processes and attributes; 2 to assess the full ecological damage caused by wetland quality decline or loss; 3 to assess the human welfare significance of such wetland changes, via a determination of the changes in the composition of the wetlands, provision of goods and services and consequent impacts on the well-being of humans who derive use or non-use benefits from such a provision; 4 to formulate practicable indicators of environmental change and sustainable utilisation of wetlands (within the D-P-S-I-R framework); 5 to carry out evaluation analysis (via a range of methods and techniques, including systems analysis) of alternative wetland change scenarios; 6 to assess alternative wetland conversion, development and conservation management policies; 7 to present resource managers and policy makers with the relevant policy response options. A number of important aspects of any economic assessment of a wetland are presented in Box 26.1. The techniques, which can be employed to value wetland functions, and each of the subsequent points in Box 26.1, are considered in more detail below. The fundamental scarcity of resources is the underlying cause for the loss of so many wetlands. Quantifying the benefits from continued wetland functioning in a way that makes them comparable with the returns derived from alternative uses of wetlands can strengthen the case for conserving wetlands, and can improve social decision-making. Cost-benefit analysis, based on the economic efficiency criterion, offers one method to aid decision makers in this context. If wetlands perform many functions and are potentially so valuable, a reasonable question would be, ‘why have these values been ignored and wetland losses and degradation allowed to continue?’ Accepting that some degree of conversion might well have been in society’s best interests where the returns from competing land use are high, wetlands have often been lost to activities resulting in only limited benefits or, on occasion, even costs to society. This is the result of what Turner and Jones (1991) refer to as

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Economic Evaluation of Wetlands

Box 26.1

Economic assessment of wetlands

Problem orientation Any analysis should take account of the prevailing political economy context, equity issues and possible ‘stakeholders’. Data inadequacies must be acknowledged and recommendations made conditional upon these limitations. Typology A useful common terminology regards functions as relationships within and between natural systems; uses refer to use, potential use, and non-use interactions between human and natural systems; and values refer to assessment of human preferences for a range of natural or nonnatural ‘objects’ and attributes. Thresholds and wetland change scenarios Thresholds relate to the scale and frequency of impacts. Their occurrence can be presented in a simple three part classification: no discernible effects; discernible effects; discernible effects that influence economic welfare. Economic valuation There are three broad approaches to a valuation exercise, the use of each depending on the type

inter-related market and intervention failures. These derive from a fundamental failure of information, or a lack of understanding of the multitude of values that may be associated with wetlands. Costanza et al. (1997) conducted a valuation of the functions of ecosystems in terms of services to humans, either directly or indirectly, through maintenance of the environment. Measurements included regulation of the water and nutrient cycles, regulation of climate, treatment of wastes, the regulation of and protection from erosion and weather and the production of food and resources. The total value of ecosystem services came to US$33 trillion, which, it was pointed out by Balmford et al. (2002), is approximately the

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of threat and the potential means of conservation: impact assessment; partial analysis and total valuation. For each function or impact, a number of techniques exist for attributing economic value to environmental benefits. Systems analysis and multi-criteria evaluation methods can complement economic cost-benefit analysis. Scale The catchment should be the minimum spatial unit for assessing ecological variables, with possible zonation within this. In terms of benefit estimation, the minimum scale is determined by the relevant population affected by any impacts. The temporal scale of analysis is also fundamentally important. Transferability (spatial and temporal) Transferring scientific results and economic benefits is problematic. Accuracy of benefit transfer may be improved if it is based on scientific variables divided into separate components depending on processes, functions and ‘state variables’.

same as total global gross national product (GNP). However, aggregate (global scale) estimates of ecosystems value are problematic, given the fact that only ‘marginal’ values are consistent with conventional decision-aiding tools such as economic cost-benefit analysis (Turner et al. 2003b). Nevertheless, in general, valuation data provide prima facie support for the hypothesis that net ecosystem service value diminishes with biodiversity and ecosystem loss (Balmford et al. 2002). A review of the costs and benefits of conversion of wild habitats to other uses, for example woodland into logs, a mangrove into aquaculture, found that although the benefits often accrued to the developer rather than the local or wider

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community, as had been the case prior to conversion, the benefits of investment in the conservation of ‘wild’ ecosystems were substantially greater than the costs (Balmford et al. 2002). The cost-benefit criterion may need to be modified as policy makers introduce, or respond to, concerns other than economic efficiency, that is, equity concerns, employment concerns, zeronet loss biodiversity conservation concerns and so on. Further, governments have now formally adopted the sustainable development policy objective, and these concerns may require the deployment of multi-criteria decision analysis methods designed to aid policy makers in policy conflict and goals trade-off situations.

SUS T AIN AB L E E CO N O M IC D E V E L OP M E N T AS A P OL ICY O B JE CT IV E Under the sustainability principle there is a requirement for the sustainable management of environmental resources, whether in their pristine state or through sympathetic utilisation, to ensure that the legacy of our current activities does not impose an excessive burden on future generations. Sustainability essentially requires that the stock of capital available in future is equivalent to that available at present. It has been suggested that it is ‘large-scale complex functioning ecologies’ that ought to form part of the intergenerational transfer of resources (Cumberland 1991). Since wetlands are complex multi-functional systems they are likely to be more beneficial if conserved as integrated ecosystems rather than in terms of their individual component parts. Wetlands mitigation in the US (Marsh et al. 1996) can be considered a specific form of imposing a ‘strong sustainability’ constraint, requiring that the loss of a wetland be compensated with an alternative wetland of equal physical quality, that is, non-declining stock of natural capital. In Europe, the Habitats Directive has a similar ‘no net loss’ principle encompassed within it, as far as designated Natura 2000 sites are concerned. This is in contrast to a ‘weak sustainability’ constraint

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in which the total stock of capital, whether manufactured or natural, be maintained.

R ISK , UN CER T AIN T Y AN D T HR ESHOLD EFFECT S Risk and uncertainty will be associated with both the physical outcomes associated with future environmental change and their economic consequences. Assessing the possible outcomes and the likelihood of perturbations to highly complex wetland ecosystems will inevitably be subject to uncertainty. However, assessment will be necessary if any economic valuation is to be considered. A range of possible impacts deriving from potential management actions – for instance a standard conserve or develop scenario – needs to be identified, the relevant physical effects quantified and probabilities attached to each. A particularly important aspect relating to the uncertainty of physical effects is the existence of thresholds beyond which disproportional and possibly irreversible effects occur. These will be important in an economic sense owing to both the disproportional extent of the impact and the inability to reverse the consequences in the future. Further discussion of the risk and uncertainty issue relating to system dynamics is covered below.

CAT EGOR IES OF ECON OMIC V ALUE In instrumentally valuing a resource such as a wetland, the Total Economic Value (TEV) can be usefully broken down into a number of categories. These are illustrated in Figure 26.4. The initial distinction is between use value and nonuse value. Use value involves some interaction with the resource, either directly or indirectly. Indirect use value derives from services provided by the wetland, for example the removal of nutrients. Direct use value, on the other hand, involves interaction with the wetland itself rather than via the services it provides. It may be consumptive use, such as the harvesting of reeds or fish, or it may be non-consumptive as

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Aesthetic/ educational use value

Distant use value

Nonconsumptive use value

Use value

Consumptive use value

Total wetland ecosystem value

Direct use value

Indirect use value Total economic value

Option value

Quasi-option value

Existence value

Bequest value

Non-use value

Philanthropic value

Fig. 26.4 Components of economic value.

with some recreational and educational activities. There is also the possibility of deriving value from ‘distant use’ through media such as television or magazines. Another category is that of option value, in which an individual derives benefit from ensuring that a resource will be available for use in

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the future. Quasi-option value is associated with the potential benefits of awaiting improved information before giving up the option to preserve a resource for future use (Freeman 1993). An example of an option value is in bio-prospecting. Non-use value is associated with benefits derived simply from the knowledge that a

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resource, such as an individual species or an entire wetland, is maintained. It is, by definition, not associated with any use of the resource or tangible benefit derived from it. Non-use value is closely linked to ethical concerns. It can be split into three basic components: existence value, derived from knowing that some feature of the environment continues to exist, whether or not this might also benefit others; bequest value, associated with the knowledge that a resource will be passed on to descendants; and philanthropic value, associated with the satisfaction from ensuring resources are available to contemporaries of the current generation. As Figure 26.4 illustrates, Total Economic Value is itself regarded as a part of the overall ‘Total Wetland Ecosystem Value’. Recent advances in the development of ecological economic models and theory point to another dimension of total environmental value: the value of the overall system itself. Here the economy and the environment are linked in a process of co-evolution, with the dynamics of the jointly determined system characterised by discontinuous change around poorly understood critical threshold values. Under the stress and shock of change, the joint systems exhibit resilience, that is, the ability of the system to maintain its self-organisation while suffering stress and shock. This resilience capacity is related to the overall system configuration and stability properties. The adoption of a systems perspective serves to re-emphasise the obvious but fundamental point that economic systems are underpinned by ecological systems and not vice versa. Biophysical systems react to economic activity that exploits environmental assets (extraction, harvesting, waste disposal and non-consumptive users). Feedbacks then occur which influence economic and social relationships. ‘Co-evolution’ is thus a crucial concept (Common and Perrings 1992). The ‘integrity’ of an ecosystem is more than its capacity to maintain autonomous functioning (its health); it also relates to the retention

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of ‘total diversity’ that is, the species and interrelationships that have survived over time at the landscape level. A number of ecological services and functions can be valued in economic terms, while others cannot because of uncertainty and complexity conditions. Private economic values may not capture the full contribution of component species and processes to the aggregate life-support functions provided by ecosystems (Gren et al. 1994). Furthermore, some ecologists argue that some of the underlying structure and functions of ecological systems, which are prior to the ecological production functions, cannot be taken into account in terms of economic values. Total Economic Value will therefore underestimate the true value of ecosystems. The prior value of the ecosystem structure has been called ‘primary value’ (PV) and consists of the system characteristics upon which all ecological functions depend (Turner and Pearce 1993). Their value arises in the sense that they produce functions that have value (secondary value). The secondary functions and values depend on the continued ‘health’, that is existence, operation and maintenance, of the ecosystem as a whole. The primary value notion is related to the fact that the system holds everything together (and is thus also referred to as a ‘glue’ value) and as such has, in principle, instrumental value. Thus, the Total Value of the ecosystem exceeds the sum of the values of the individual functions. It can also be argued that a healthy ecosystem contains an ecological redundancy capacity and there is thus an ‘insurance’ value in maintaining the system at some ‘critical’ size in order to combat stress and shocks over time. We now are in a position to bring together all the elements in the total environmental value (TV) framework. These are laid out in Table 26.2 and Figure 26.5, with TEV as a part of TV, and PV and quasi-option value as separate and non-commensurate value components. Non-anthropocentric intrinsic value is best viewed as a completely separate notion not commensurate with any of the other components.

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Economic Evaluation of Wetlands Table 26.2 A general value typology. (Adapted from Hargrove, 1992.) Anthropocentric Instrumental value

Intrinsic value

Non-anthropocentric Instrumental value

Intrinsic value

‘Total Economic Value’ = use value + non-use value. The non-use category is bounded by the existence value concept which has been the subject of much debate. Existence value may therefore encompass some or all of the following motivations

interpersonal altruism (philanthropic motivation and value), resource conservation to ensure availability for others; vicarious use value linked to self-interested altruism and the ‘warm glow’ effect of purchased moral satisfaction intergenerational altruism (bequest motivation and value), resource conservation to ensure availability for future generations stewardship motivation, human responsibilities for resource If existence value is defined conservation on behalf of all nature to include stewardship and Q-altruism, motivation based on the ‘Q-altruism’ it will overlap belief that non-human resources into the next category have rights and interests and, outlined below as far as possible, should be left undisturbed

This value category is linked to anthropocentrism that recognises a range of values extending beyond instrumental values. It is culturally dependent

Value attribution is to entities which have a ‘sake’ or ‘goods of their own’, and ‘instrumentally use other parts of nature for their own intrinsic ends ...’. It remains an anthropocentrically related concept because it is still a human value that is ascribing intrinsic value to non-human nature (‘Q-altruism’)

Entities are assumed to have sakes or goods of their own, independent of human interests.

Also encompasses the good of collective entities, e.g. ecosystems, in a way that is not irreducible to that of its members. This category may not demand moral consideration as far as humans are concerned

Viewed in an objective sense, i.e. ‘inherent worth’ in nature, the value that an object possesses independently of the valuation by humans

A meta-ethical claim, and usually involves the search for rules or trump cards with which to constrain anthropocentric instrumental values and policy. It is therefore entirely separate from any human-related environmental value component

     

C O M BIN IN G E CON OM ICS AND W E T L AN DS S CIE N CE The links between wetland functions and wetland values are illustrated in Figure 26.2. What is clear is that wetland uses, or the output of physical products or services, form the essential link between wetland ecology or functioning and wetland economics or values. It is important, therefore, to identify how particular functions

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might be of use, rather than simply the degree to which the function is being performed. The extent of demand for the products or services provided, or the effective ‘market’, needs to be evaluated if the full extent of economic value is to be assessed. Given the general typology for the assessment of wetland benefits provided in Figure 26.2, the first step is to compile a complete list of all the characteristics of a wetland. Characteristics

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Use value ? DUV + IVV + OV +

TEV =

Non-use value ? BV

+

Anthropocentric instrumental value

TV* =

PV

and

TEV

and

EV

Anthropocentric intrinsic value

QOV

IV

Anthropocentric instrumental value

Non-anthropocentric intrinsic value

Non-anthropocentric instrumental value

TV TEV DUV IUV OV QOV EV PV IV BV

= = = = = = = = = =

Total Environmental Value Total Economic Value Direct Use Value Indirect Use Value Option Value (including Bequest Value) Quasi Option Value Existence Value Primary Value Intrinsic Value Bequest Value

*Note: the separate components of TV are not additive, they are different dimensions of value. Fig. 26.5 Total Environmental Value and Total Economic Value.

are those properties that describe the area in the simplest and most objective terms, including the biological, chemical and physical features such as size, shape, depth, climate, species present, vegetation structure and the natural processes occurring there. From an anthropocentric viewpoint, all ecosystems can be classified in terms of their structural and functional aspects (Westman 1985; Turner 1988). Ecosystem structure is defined as the tangible items such as plants, animals, soil, air and water that comprise it, that is, what it has. It provides humans with goods or products, which

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involves some direct utilisation of one or more of the characteristics of a wetland. By contrast, ecosystem functions or processes are encompassed by the dynamics in the system, that is, what it does. The processes are subsequently responsible for the ecologically related services – life support services, such as assimilation of pollutants, cycling of nutrients and maintenance of the balance of gases in the air. The task of evaluating the structure and function of an ecosystem implies that we know fully what the ecosystem does and what that worth is to us. The worth of ecosystem structure

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Economic Evaluation of Wetlands Table 26.3 Classification of benefits of wetlands as services, goods and attributes. SERVICES flood control prevention of saline intrusion storm protection or windbreak sediment removal toxicant removal nutrient removal groundwater recharge groundwater discharge erosion control wildlife habitat fish habitat toxicant export shoreline stabilisation microclimate stabilisation macroclimate stabilisation biological diversity provision cultural value provision historic value provision aesthetic value provision wilderness value provision

GOODS forest resources agriculture resources wildlife resources forage resources fisheries mineral resources water transport water supply recreation and tourism aquaculture research site education site fertiliser production energy production

is generally more easily appreciated than that of ecosystem functions. To evaluate functions pushes present ecological knowledge beyond its bounds. Even ecosystem structure is incompletely known. Much research on ecosystem structure and functioning is still needed. An assessment of the complete range of benefits at a wetland site using a standard classification of benefits, as listed in Table 26.3, is an essential step before the overall value of the wetland can be derived. It is evident that there are strong linkages between the different types of benefit, and the need to ensure against double counting cannot be overstated.

SYSTE M S AN AL Y S IS AN D D Y N AMICS: E NV IR ON M E N T AL CHAN GE S CE N AR IOS Analysts should not assume that all current economic conditions will hold in the future.

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615

For instance, land use changes might be predicted for the future, perhaps due to imminent regulation or long-term trends. This might affect, for example, the quantity of nitrogen in runoff and thereby the value of the wetland as a nitrogen sink. Behaviour of individuals could also adapt to change in wetland functioning, for instance with farmers changing cropping patterns as a result of increased flooding, rather than forgoing landuse or yields altogether. These changes need to be incorporated into the analysis since they can influence projected benefits and, hence, the net present value associated with maintaining wetland functions. Uncertainty over the correct value for economic variables employed, and future environmental and socio-economic changes and trends can be addressed by employing a sensitivity analysis or scenario analysis. As Costanza (1994) points out, ‘most important environmental problems suffer from true uncertainty, not merely risk’. In an economic sense, such pure uncertainty can be considered as ‘social uncertainty’ or ‘natural uncertainty’ (Bishop 1978). Social uncertainty derives from factors such as future incomes and technology, which will influence whether or not a resource is regarded as valuable in the future. Natural uncertainty is associated with our imperfect knowledge of the environment and whether there are unknown features of it that may yet prove to be of value. This might be particularly relevant to wetland ecosystems where the multitude of functions that are being performed have historically been unappreciated. One practical means of dealing with such complete uncertainty is to complement a costbenefit criterion based purely upon monetary valuation, with a Safe Minimum Standards (SMS) decision rule (Ciriacy-Wantrup 1952; Bishop 1978; Crowards 1996). This recommends that, when an impact on the environment from a development threatens to breach an irreversible threshold, the conservation option be adopted unless the costs of forgoing the development are regarded as ‘unacceptable’. It is based on a principle of minimising the maximum possible loss. Given our absolute uncertainty regarding

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the benefits in the future that might be derived from the threatened resource, our maximum possible loss must be associated with loss of that resource. Clearly, the critical factor in SMS is defining what is an ‘unacceptable’ sacrifice of present benefits for the sake of possible future losses. The concept of Safe Minimum Standards is usually applied to endangered species. In this manner, it may well be applicable to a number of wetlands given their role in supporting a variety of threatened species. However, it could equally well apply to irreversible impacts threatening wetland ecosystems as a whole. One complication is to identify what is a truly irreversible change in the ecosystem, since any change that can be reversed in the future will not necessarily entail the maximum possible costs. It will also be necessary to determine whether or not thresholds in current wetland functioning exist, and whether these may be threatened by proposed developments.

P RA C T ICAL IS S U E S O F E CO N O MIC V AL UAT IO N Scale It is important to determine initially what the scale of assessment is going to be. This may involve impact analysis of a limited number of affected variables associated with an isolated external impact. Where more general changes, such as alternative uses of the wetland, are being considered, partial analysis of a number of integrated parameters may be required. A total valuation will generally entail considerable effort but may be appropriate where, (i) the wetland as a whole is threatened, and (ii) the benefits deriving from an alternative use outweigh the benefits estimated from only a partial estimation of wetland value. The geographical scale of assessment will be important. The relevant population for an economic assessment will depend in part on the type of function that is being valued. Direct use values

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will generally involve some contact with the wetland itself, although individuals may travel considerable distances in order to make use of the wetland. Indirect use values may be sitespecific in terms of those who benefit; non-use values are likely to be derived over a wide geographical range, but are likely to be subject to ‘distance decay’ away from the wetland site. Temporal scale, in combination with the rate of discount applied, will influence the present value of benefits attributed to wetland functions. Calculating expected future costs and benefits involves estimating future demand for the wetland’s functions. This will necessarily be unknown, but assessing likely scenarios and applying sensitivity analysis can provide a range of possible values. Problem orientation, data requirements and benefits transfer The current economic, political and cultural climate within which the study is being carried out should be assessed. This will influence estimates of the future trend scenario, other alternative scenarios based on alternative management, and policy responses and essential parameters, such as the rate of discount to be applied. The pattern of resource ownership might also determine the extent to which market prices exist for goods and services, and indicate the importance of wider social goals other than economic efficiency. Once the scale of assessment and its institutional setting has been determined, the information requirements can be analysed. Inevitably, not all data will be readily available and budgetary constraints are likely to limit extensive collection of primary data. Where data are limited this should be acknowledged, the response to this limitation outlined, and any results and recommendations that transpire should explicitly be made conditional on these limitations. In order to estimate benefits, when given limited funds and a relatively short time scale, it may be possible to transfer data from other studies as a rough guide to appropriate values. This technique of ‘benefits

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Economic Evaluation of Wetlands transfer’ is, however, fraught with difficulties and subject to a number of caveats. Criteria for transferring benefits between sites (based on Boyle and Bergstrom 1992) as: 1 valuations should be of the same goods or services in each case; 2 relevant populations should be similar; 3 the assignment of property rights concerning the wetland function under consideration should be the same. Three approaches to benefits transfer are outlined by Pearce et al. (1994): 1 directly transferring mean unit values; 2 transferring unit values adjusted to suit the current study; and 3 transfer of a benefit function from which unit values can be derived. A major drawback of the direct transfer of values is that no two situations will be identical and the criteria outlined above are unlikely to be met. Values will need to be adjusted when there are differences in socio-economic characteristics of households, differences in the availability of substitute or complementary goods or services, and differences in the policy setting and problem orientation. The transferring of benefit functions is likely to result in better approximation of appropriate values, but is more involved than the other two approaches. Problems common to all methods of benefits transfer remain: the requirement for good quality studies of similar situations, the considerable potential for changes in characteristics between different time periods; and the inability to evaluate novel changes. Meta analysis and wetland functions Since the beginning of the 1990s, meta-analysis has been playing an increasingly important role in environmental economics research (van den Bergh and Button 1997). Meta-analysis is the statistical evaluation of the summary findings of empirical studies, helping to extract information from large masses of data in order to quantify a more comprehensive assessment. It enables researchers to explain differences in outcomes

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found in single studies on the basis of differences in underlying assumptions, standards of design,and measurement. Meta-analysis enables researchers to identify criteria for valid environmental value transfer1 or to test the convergent validity of value estimates. In the first case, the data set is entirely used to determine the factors that help to explain variances in valuation outcomes. In the second case the data set can be split, for example into two parts, one of which is used for the first purpose and another to test whether the value estimates based on the significant factors fall within the confidence interval of the other half’s estimates. The criteria for selecting studies for environmental value transfer suggested in the literature focus on the environmental goods involved, the sites in which the goods are found, the stakeholders and the study quality (Desvousges et al. 1992). However, very little published evidence exists of studies that test the validity of environmental value transfer. Moreover, in the few studies that have been carried out, the transfer errors are substantial (Brouwer 1998). Valuation techniques A range of valuation techniques exists for assessing the economic value of the functions performed by wetlands, and these are detailed in Table 26.4. Many wetland functions result in goods and services that are not traded in markets and therefore remain un-priced. It is then necessary to assess the relative economic worth of these goods or services using non-market valuation techniques. More detailed information on the underlying theory and practical implementation of these techniques can be found in a number of general texts including Braden and Kolstad (1991), Bromley (1995), Dixon and Hufschmidt (1986), Freeman (1993), Hanley and Spash (1993), Pearce et al. (1994), Randall (1987), Turner (1993), and Turner and Adger (1996). An important distinction to make is between those valuation techniques that estimate benefits directly and those that estimate costs as a proxy for

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Table 26.4 Valuation methodologies relating to wetland functions. Valuation method

Description

Direct use values

Indirect use values

Market Analysis

Where market prices of outputs (and inputs) are available. Marginal productivity net of human effort or cost. Could approximate with market price of close substitute. Requires shadow pricing. Change in net return from marketed goods: a form of (dose–response) market analysis. Wetlands treated as one input into the production of other goods: based on ecological linkages and market analysis. Public investment, for instance via land purchase or monetary incentives, as a surrogate for market transactions. Derive an implicit price for an environmental good from analysis of goods for which markets exist and which incorporate particular environmental characteristics. Costs incurred in reaching a recreation site as a proxy for the value of recreation. Expenses differ between sites (or for the same site over time) with different environmental attributes. Construction of a hypothetical market by direct surveying of a sample of individuals and aggregation to encompass the relevant population. Problems of potential biases. The costs that would be incurred if the wetland function were not present; e.g. flood prevention. Costs incurred in mitigating the effects of reduced environmental quality. Represents a minimum value for the environmental function. Expenditures involved in relocation of affected agents or facilities: a particular form of defensive expenditure. Potential expenditures incurred in replacing the function that is lost; for instance by the use of substitute facilities or ‘shadow projects’. Costs of returning the degraded wetland to its original state. A total value approach; important ecological, temporal and cultural dimensions

√

√

√

√

(Productivity Losses) (Production Functions) (Public Pricing)

Hedonic Price Method (HPM)

Travel Cost Method (TCM)

Contingent Valuation (CVM)

Damage Costs Avoided Defensive Expenditures

(Relocation Costs) Replacement/Substitute Costs

Restoration Costs

benefits. For instance, estimating Damage Costs Avoided, Defensive Expenditures, Replacement/ Substitute Costs or Restoration Costs as part of an economic valuation exercise suggests that the costs are a reasonable approximation of the benefits that society attributes to the resources in question. Where it can be shown that, (i) replacement or repair will provide a perfect substitute for the original function, and (ii) the costs of doing so are less than the benefits derived from

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Non-use values

√ √

√

√

√

√

√

√

√

√

√

√ √

√ √

√

√

√

√

√

this function, the costs do indeed represent the economic value associated with that function. Where market prices exist for resources, these may have to be adjusted to provide social or shadow prices as explained above, but otherwise they are likely to provide a relatively simple means of assessing economic value. Approaches related to market analysis include the assessment of productivity losses that can be attributed to changes in the wetland, and the theoretical

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Economic Evaluation of Wetlands incorporation of the wetland as just one input into the production function of other goods and services. Investment by public (especially government) bodies in conserving wetlands may represent a surrogate for aggregated individual willingness to pay and hence social value. In the absence of market prices, two theoretically valid benefit estimation techniques would be hedonic pricing or the travel cost method. However, these are based on preferences being ‘revealed’ through observable behaviour, and are restricted in their application to where a functioning market exists, such as that for property in the case of hedonic pricing, or where travel to the site is a prerequisite to deriving benefit, such as with recreational visits, in the travel cost method. Contingent valuation, based on surveys that elicit ‘stated preferences’, has the potential to value benefits in all situations, including non-use benefits that are not associated with any observable behaviour. The legitimacy of contingent valuation methods and results is still contested, especially in the context of non-use values, and conducting a contingent valuation survey can sometimes be a lengthy and resource-intensive exercise. Aggregation and double counting If each output provided by the wetland is identified separately, and then attributed to underlying functions, there is the likelihood that benefits will be double counted. Benefits might therefore have to be explicitly allocated between functions. For instance, Barbier (1994) notes that if the nutrient retention function is integral to the maintenance of biodiversity, then if both functions are valued separately and aggregated this would double count the nutrient retention which is already ‘captured’ in the biodiversity value. Double counting will be particularly important with partial analysis and total valuation of a wetland, although some approximations to total valuation do not encounter this problem. Studies that attempt to value the wetland as a whole based on an aggregation of separate values will tend to include a certain number of functions, although

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most do not usually claim to cover all possible benefits associated with the wetland. Opportunity costs A number of wetlands studies have considered the opportunity costs of wetland valuation, such as forgone agricultural production or other development benefits, as an alternative to estimating the value of the wetland itself. For instance, Turner et al. (1983) estimate the opportunity cost of preserving wetlands in the Norfolk Broads of the UK, as the potential net returns to arable farming on the same land. The returns to this alternative use were found to be negative, and there was therefore no opportunity cost to preserving the wetlands and, in this instance, no need to estimate their economic value. Where the benefits of an alternative use are positive, qualitative assessment of the range of functions performed by a wetland and the values that could be associated with these, without actually quantifying them, may suggest that preservation is likely to be the preferred option (for instance, Bucher and Huszsr 1995). It should be noted, however, that opportunity cost analysis does not represent a valuation of the wetland, but simply a valuation of the alternative that threatens wetland functioning.

EV ALUAT ION FR OM A DY N AMIC SY ST EMS PER SPECT IV E Quantifying the wetland function will not in itself be sufficient. The essence of an overall socioeconomic evaluation is to determine how society is affected by the functions a wetland might perform – the function itself is not intrinsically valuable. It will therefore be necessary to assess features of anthropogenic regimes, especially downstream of the wetland, and how these respond to changes in wetland functioning. This is in contrast to the physical procedures in which much of the emphasis is on the influence that factors upstream can have on the wetland. Furthermore, economic analysis is not limited to

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areas with functional linkages to the wetland, but is generally more concerned with the economic region of influence and the range of relevant stakeholder interests and positions. This may conform somewhat to patterns of the local physical environment but is by no means determined by it. A number of initial points that might be considered include: 1 Proximity of the wetland, and of the result of the functions it performs, to human settlement; 2 Accessibility of the wetland to humans; 3 Land use in the vicinity of and downstream from the wetland; 4 Predominant local industries such as farming, manufacturing, mining, tourism, forestry – not limited to the catchment, but the local region; 5 Size and distinctiveness of the wetland in comparison with other nearby wetlands; 6 Likely changes in human factors in the catchment in the foreseeable future, such as urban expansion, change in farming practices, increasing nature conservation, road building, river management etc; 7 The existence of nature conservation schemes within or near the wetland; 8 Effluent discharge or nutrient seepage into a river upstream or downstream of the wetland; 9 Known or historical problems with the water environment, such as pollution of the river, or flooding episodes. Since economic analysis is resource intensive, it will be useful to have a screening process to determine the key functions of value before attempting a valuation study. Some initial indicators of significant wetland value could include its designation as a conservation area, extensive harvesting of products, extraction of water, direct input of effluent, ongoing and historical river or flood management schemes and recreational usage. It will also be the case that economic efficiency, although important, will not be the only decision criterion of significance to resource managers and policy makers. A number of so-called multi-criteria

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decision-support analysis methods (Nijkamp 1989; Janssen 1992) have been developed in order to illuminate policy trade-offs and aid decision making in contexts where a range of, often competing, policy criteria are considered to be socially and politically relevant.

IN T EGR AT ED ECOLOGICAL– ECON OMIC MODELLIN G OF WET LAN DS A proper interdisciplinary analysis of wetland functions, uses and management must be based on an overall systems perspective and approach. Figure 26.6 summarises the main components of such a methodology and illustrates how the different disciplines can contribute in a coordinated and overlapping fashion. Integrated modelling comes in two forms. One strives towards a single model, while the other employs a system of heuristically connected submodels. Coupling wetland ecology and wetland economics within one integrated model inevitably involves compromises and simplifications. A number of approaches to integrated modelling exist, based on generalised input–output models, nonlinear dynamic systems models, optimisation models, land use models linked to geographical information systems (GIS) and mixed models. Important elements for integration are connected scenarios, models and indicators, and the arrangement of consistency among units, spatial demarcations and spatial aggregation of information in various submodels. An overview of integrated modelling approaches and applications is given in van den Bergh (1996).

T OWAR DS AN IN T EGR AT ED FR AMEWOR K FOR WET LAN D ECOSY ST EM IN DICAT OR S The main objective here is to develop an integrated framework for environmental indicators

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Biogeochemical systems Models and analysis: process, functions and ecosystem dynamics

Information systems e.g. GIS and related methods

Socio-economic systems Models and analysis: evaluation methods, scenario formulation and policy analysis

Fig. 26.6 Interdisciplinary resource assessment approach.

to monitor and evaluate the impact of human intervention (e.g. pollution, conversion, management) on wetland ecosystems in a way which can inform policy and decision-making. The maintenance of ecosystem integrity is the core objective which underpins the indicator sets which will be presented and discussed. The concept of ecosystem integrity has recently gained in popularity and has been defined as the maintenance of system components, interactions among them, and the resultant behaviour or dynamic of the system (King 1993). In the literature, other ‘sustainability’ concepts have been proposed as well, such as ‘ecosystem health’ (Costanza et al. 1992) or ‘ecosystem resilience’ (Holling 1986). The former has been defined as a system free of distress syndrome, while the latter refers to a system’s ability to maintain its structure and pattern of behaviour in the presence of stress. Ecosystem integrity is compatible with these other concepts in that they all refer to, implicitly or explicitly, a certain minimum structural system composition required for the overall functioning of ecosystems. The loss of a single system component, such as the loss of a single species or population (resulting in a change of ecosystem composition or

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structure) or a change in interaction (processes), does not necessarily imply loss of system integrity. Many systems, including ecological systems, appear to be resilient to alteration of structure. Whole system function is maintained despite the structural change. The complex land–water interactions found in wetlands and their openness necessitate a management approach that accounts for these two key characteristics in a coherent and consistent way. The focus will be on existing ecological, biological, chemical and hydrological indicators to monitor and evaluate wetland management. However, in view of the use of wetlands and the values attached to them by human beings, relevant social and economic aspects cannot be excluded if the ultimate aim is to present an integrated set of indicators for sustainable resource management. In this sense, the presented environmental indicator sets are an intermediate step towards such an integrated information format. Two of the major problems surrounding indicators are the scientific uncertainty over whether they actually measure features of the environment that are of interest, and that they change in some meaningful way with respect to

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environmental change (Norris and Norris 1995). Ideally, one indicator could be constructed that can serve as an index of the structure and functioning of an entire ecosystem, but given the complexity of natural systems, this will rarely be the case. King (1993) argues that the description of a system simultaneously involves both structure and function. When attempting to describe a system, one should ask: what are the components, how are they connected and how do they operate together? According to King, system integrity thus implies the integrity of both system structure and function, as well as maintenance of system components, interactions among them and the resultant behaviour or dynamic of the system. Given the emphasis on both system components and the interactions between them, this implies that the loss of system integrity has to be evaluated in the light of both structure and function or processes. In the context of biological diversity, Franklin et al. (1981) distinguished a third criterion to identify and describe ecosystems. Ecosystem composition, in their terms, refers to the identity and variety of elements in a system, whereas structure refers to the physical organisation of a system. An organising framework for wetland system indicators is summarised in Table 26.5. The effect of pressures exerted by human activities on ecosystems can be measured by defining the relevant indicator spheres for ecosystem structure, composition and function. The sustainability of natural resource management or the impacts of environmental pressures on ecosystem integrity can subsequently be assessed in the following two ways: First, indicators of environmental change can be related to ecological, biogeochemical and hydrological benchmarks or ‘sustainability rules’ based on natural scientific models. For example, emission loads in an ecosystem can be compared with the assimilative capacity of a specific ecosystem. Second, ecosystem integrity can be assessed by investigating the impact of environmental pressures on structural, compositional and functional ecosystem changes and the impact of these

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Table 26.5 Organising framework for wetland ecosystem indicators. Wetland domain

Ecosystem component

Indicator sphere

Landscape

Structure Composition Function

Topology (spatial connectedness) Land form and cover Land use interactions

Water regime Structure Composition Function

Hydrology Biogeochemical water properties Water constituents flux

Biodiversity

Food web trophic structure Keystone species and umbrella species Energy transfer between trophic levels

Structure Composition Function

changes on each other. This will often be a qualitative analysis. For example, human induced stress may alter ecosystem structure or composition. The impact of this change in ecosystem structure or composition on general ecosystem functioning has to be investigated to assess ecosystem integrity. Although agencies seem to prefer to promulgate and enforce regulations based on quantitative criteria, descriptions of qualitative changes in, for example, community structure, are often the best indicators of ecological disruption (Noss 1990). In practice, qualitative descriptions of the intermediate changes or transitions between ecosystem states and ecosystem functions may sometimes prove the only way (or a very important complement to numerical indicators) to assess the extent to which wetland management restores, maintains or enhances the integrity of an ecosystem. Given the unique interactions between water and land in wetland ecosystems, these two environmental media are used as the main indicator ‘domains’. Biodiversity is the third domain. Based on the specific indicators found in the literature, indicator sets have been compiled for each of these domains following the conceptual breakdown of ecosystems into structure, composition and

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Economic Evaluation of Wetlands function components (see Table 26.5). Although this breakdown seems to suggest that these aspects will be looked at independently, it is important to emphasise that it is the interdependency and compatibility within and between ecosystem components and wetland domains, across different scales and hierarchies, that is of primary interest. The impacts of human activities or management intervention on wetland ecosystems can be assessed only if these three indicator spheres are considered together. Brouwer et al. (1998) discuss the three wetland domains and their indicator spheres in more detail.

CON CL US ION S At the wetland ecosystem level, it has proved possible to confirm a coherent set of links between wetland boundary conditions, structure, processes and functions, as well as the consequent outputs of goods and services from which humans derive both use and non-use value. The precise quantification and valuation of multiple wetland functions and outputs is not, however, straightforward. Function overlap (a double counting problem) and conflict (one use precluding another use) and the delineation of the overall ecological system integrity (‘infrastructure’ or ‘primary’ value of wetlands) present a number of complications. Nevertheless, this chapter has confirmed the importance of the ‘functional’ approach and the related policy objective of ‘maintaining functional diversity’ at the catchment or landscape scale. At the landscape scale, the D-P-S-I-R auditing framework has also proved to be a very useful device. It allows us to coherently address a range of issues resulting from regional and local environmental pressures and resource use conflicts.

E N DN OT E 1

The term ‘environmental value transfer’ is used here instead of the popular term ‘benefits transfer’, because economists can also measure the benefits foregone, which turns the estimated values into

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costs instead of benefits. Willingness to pay (WTP) is the conventional economic approach to measure environmental values in money and hence make them commensurable with other market values (costs and benefits) associated with decisions that have been made, are made or have to be made in the face of limited human and natural resources.

R EFER EN CES Balmford A., Bruner A., Cooper P., Costanza R., Farber S., Green R.E., Jenkins M., Jefferiss P., Jessamy V., Madden J., et al. 2002. Economic reasons for conserving wild nature. Science 297, 950–953. Barbier E.B. 1994. Valuing environmental functions: tropical wetlands. Land Economics 70(2), 155–173. Barbier E.B., Acreman M. and Knowler D. 1997. Economic Valuation of Wetlands: A Guide for Policy Makers and Planners. Ramsar Convention Bureau, Gland. Bishop R.C. 1978. Endangered species and uncertainty: the economics of a safe minimum standard. American Journal of Agricultural Economics 60, 10–18. Boyle K.J. and Bergstrom J.C. 1992. Benefit transfer studies: myths, pragmatism, and idealism. Water Resources Research 28(3), 657–663. Braden J.B. and Kolstad C.D. (editors) 1991. Measuring the Demand for Environmental Quality. NorthHolland, Amsterdam. Bromley D.W. (editor) 1995. The Handbook of Environmental Economics. Blackwell, Oxford. Brouwer R. 1998. Future Research Priorities for Valid and Reliable Environmental Value Transfer. Global Environmental Change Working Paper 98–28, Centre for Social and Economic Research on the Global Environment (CSERGE), University of East Anglia and University College London. Brouwer R., Crooks S. and Turner R.K. 1998. Towards an Integrated Framework for Wetland Ecosystem Indicators. Global Environmental Change Working Paper GEC 98–27, Centre for Social and Economic Research on the Global Environment, University of East Anglia and University College London. Bucher E.H. and Huszar P.C. 1995. Critical environmental costs of the Paraguay-Parana Waterway Project in South America. Ecological Economics 15, 3–9. Ciriacy-Wantrup S.V. 1952. Resource Conservation: Economics and Policies. University of California Press, Berkeley.

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Common M. and Perrings C. 1992. Towards and ecological economics of sustainability. Ecological Economics 6, 7–34. Costanza R., Norton B.G. and Haskell B.D. (editors) 1992 Ecosystem Health: New Goals for Environmental Management. Island Press, Washington, DC. Costanza R. 1994. Three general policies to achieve sustainability. In: Jansson A.M., Hammer M., Folke C. and Costanza R. (editors), Investing in Natural Capital: The Ecological Economics Approach to Sustainability. Island Press, Washington, DC, pp. 392–407. Costanza R., d’Arge R., De Groot R., Farber S., Grasso M., Hannon B., Limburg K., Naeem S., O’Neill R.V.O., Paruelo J., et al. 1997. The value of the world’s ecosystem services and natural capital. Nature 387, 253–260. Cowardin L.M., Carter V., Gollet F.C. and LaRoe E.T. 1979. Classification of Wetlands and Deep Water Habitats of the United States. US Fish and Wildlife Service Publication FWS/OBS-79/31. Washington, DC. Crowards T.M. 1996. Addressing Uncertainty in Project Evaluation: The Costs and Benefits of Safe Minimum Standards. Global Environmental Change Working Paper GEC 96-04, Centre for Social and Economic Research on the Global Environment (CSERGE), University of East Anglia and University College, London. Cumberland J.H. 1991. Intergenerational transfers and ecological sustainability. In: Costanza R. (editor), Ecological Economics: The Science and Management of Sustainability. Columbia University Press, New York, pp. 355–366. Desvousges W.H., Naughton M.C. and Parsons G.R. 1992. Benefit transfer: conceptual problems in estimating water quality benefits using existing studies. Water Resources Research 28(3), 675–683. Dixon J.A. and Hufschmidt M.M. 1986. Economic Valuation Techniques for the Environment. John Hopkins Press, Baltimore. Dugan P.J. 1990. Wetland Conservation: A Review of Current Issues and Required Action. IUCN, Gland, Switzerland. European Environment Agency. 2000. Cloudy Crystal Balls: An Assessment of Recent European and Global Scenario and Models. Environmental Issues Series 17, European Commission, Brussels. Franklin J.F., Cromack K. and Denison W. 1981. Ecological Characteristics of Old-Growth DouglasFir Forests. USDA Forest Service General Technical

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Report PNW-118. Pacific Northwest Forest and Range Experiment Station, Portland, OR. Freeman A.M.I. 1993. The Measurement of Environmental Resources Values. Resources for the Future, Washington, DC. Gren I-M., Folke C., Turner R.K. and Bateman I. 1994. Primary and secondary values of wetland ecosystems. Environmental and Resource Economics 4, 55–74. Hanley N. and Spash C.L. 1993. Cost-Benefit Analysis and the Environment. Edward Elgar, Vermont. Holling C.S. 1986. The resilience of terrestrial ecosystems: local surprise and global change. In: Clark W.C. and Munn R.E. (editors), Sustainable Development of the Biosphere. Cambridge University Press, Cambridge, pp. 292–317. Janssen R. 1992. Multiobjective Decision Support for Environmental Management. Kluwer, Dordrecht. King A.W. 1993. Considerations of scale and hierarchy. In: Woodley S., Kay J. and Francis G. (editors), Ecological Integrity and the Management of Ecosystems. St. Lucie Press. Sponsored by Heritage Resources Centre, University of Waterloo and Canadian Parks Service, Ottawa. Larson J.S., Adamus P.R. and Clairin Jr. E.J. 1989. Functional Assessment of Freshwater Wetlands: A Manual and Training Outline. WWF and the Environmental Institute, University of Massachusetts, Amherst. Maltby E., Hogan D.V. and McInnes R.J. (editors) 1995. Functional Analysis of European Wetland Ecosystems. Final Report – Phase One. EC DGXII STEP Project. CT90–0084. Wetland Ecosystems Research Group. University of London, London. Marsh L.L., Porter D.R. and Salvesen D.A. 1996. Mitigation Banking: Theory and Practice. Island Press, Washington, DC. Nijkamp P. 1989. Multi-criteria analysis: a decision support system for sustainable environmental management. In: Archibugi F. and Nijkamp P. (editors), Economy and Ecology: Towards Sustainable Development. Kluwer, Dordrecht, pp. 203–220. Norris R.H. and Norris K.R. 1995. The need for biological assessment of water quality: Australian perspective. Australian Journal of Ecology 20, 1–6. Noss R.F. 1990. Indicators for monitoring biodiversity: a hierarchical approach. Conservation Biology 4(4), 355–364. Pearce D.W., Whittington D. and Georgiou S. 1994. Project and Policy Appraisal: Integrating Economics and Environment. OECD, Paris.

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Economic Evaluation of Wetlands Ramsar. 1971. Ramsar Convention on Wetlands of International Importance especially as Waterfowl Habitat, Ramsar, Iran 1971 (http://www.ramsar.org). Randall A. 1987. Resource Economics: An Economic Approach to Natural Resource and Environmental Policy. Wiley, New York. Scott D.A. (editor) 1989. A Directory of Asian Wetlands. IUCN, Gland, Switzerland and Cambridge, UK. Turner R.K. 1988. Wetland conservation: economics and ethics. In: Collard D., Pearce D. and Ulph D. (editors), Economics, Growth and Sustainable Development. Macmillan, London, pp. 121–159. Turner R.K. (editor) 1993. Sustainable Environmental Economics and Management: Principles and Practice. Belhaven Press, London. Turner R.K. and Adger W.N. 1996. Coastal Zone Resources Assessment Guidelines. LOICZ/R&S/ 96-4, LOICZ. IGBP, Stockholm. Turner R.K., Dent D. and Hey R.D. 1983. Valuation of the environmental impact of wetland flood protection and drainage schemes. Environment and Planning A 15, 871–888. Turner R.K. and Jones T. (editors) 1991. Wetlands, Market and Intervention Failures. Earthscan, London.

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Turner R.K., Paavola J., Cooper P., Farber S., Jesamy V. and Georgiou S. 2003b. Valuing nature: lessons learned and future research directions. Ecological Economics 46, 493–510. Turner R.K. and Pearce D.W. 1993. Sustainable economic development: economic and ethical principles. In: Barbier E.B. (editor), Economics and Ecology: New Frontiers and Sustainable Development. Chapman and Hall, London, pp. 176–194. Turner R.K., van den Bergh C.J.M. and Brouwer R. (editors) 2003a. Managing Wetlands: An Ecological Economics Approach. Edward Elgar, Cheltenham. Turner R.K., van den Bergh C.J.M., Soderqvist T., Barendregt A., van der Straaten J., Maltby E. and van Ierland E.C. 2000. Ecological–economic analysis of wetlands: scientific integration for management and policy. Ecological Economics 35, 7–23. Van den Bergh J.C.J.M. 1996. Ecological Economics and Sustainable Development: Theory, Methods and Applications. Edward Elgar, Cheltenham. Van den Bergh J.C.J.M. and Button K.J. 1997. Metaanalysis of environmental issues in regional, urban and transport economics. Urban Studies 34(5–6), 927–944. Westman R.E. 1985. Ecology, Impact Assessment and Environmental Planning. Wiley, Chichester.

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Section V Wetland Dysfunctioning: What Happens When Wetlands do not Work?

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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Introduction – How Do Wetlands Fail? K AT HERI NE C. EWEL 1,2 1USDA

2School

Forest Service, Hilo, USA of Forest Resources and Conservation, University of Florida, Gainesville, USA

IN T R O D U CT ION A wetland cannot fail unless something is expected of it, and society has indeed come to expect a great deal of wetlands within the last 50 years. Once generally regarded as wastelands and subject to ‘reclamation’, that is drainage for conversion to agriculture or urban development, wetlands now are recognised as providing an array of services ranging from maintenance of water quality and biodiversity to provision of beautiful scenery (Table 27.1). Some services, such as protection from storm tides, benefit local communities, whereas others, such as climate modification, can affect the entire globe. With several decades of observation and research behind us, we now understand that not every type of wetland can provide every type of service (Ewel et al. 1998). We also realise that any given type of wetland may provide a particular service only under a prescribed range of conditions. By the time the importance of wetlands was widely recognised, many wetlands had already been destroyed, often beyond recovery, in numerous places around the world. Pleas for saving the remaining wetlands have been based on our general understanding of the values listed in Table 27.1 Save our wetlands, we believed, and we will be saved from ourselves. Restoration of degraded wetlands, and even creation of new

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Table 27.1 Ecosystem services commonly ascribed to wetlands. Sustaining biodiversity Evolution of unique species, many of which are threatened and endangered Production of harvestable wildlife, including fish and shellfish Production of harvestable vegetation Damping variability in water resources Flood mitigation Storm tide abatement Aquifer recharge Water quality maintenance Moderating biogeochemical processes Sediment trapping Nutrient transformations Carbon storage Aesthetic value Provision of a vista Provision of wilderness and solitude

wetlands to replace those lost to development, have been strategies consistent with this goal. A broader appreciation of wetlands has resulted, although much remains to be learned. The other chapters in this section describe in detail how alterations in hydrology and nutrient enrichment in catchments (watersheds) affect wetlands (often indirectly or inadvertently), how wetlands respond to direct management in agriculture, forestry and peat extraction, and how species invasions alter wetland function. The purpose of this chapter is to describe how failure of a wetland to provide a service can be a result of our own faulty assumptions about how a wetland operates, and our unrealistic expectations about what it can do.

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U ND E R S T AN D IN G W E T L AN D S: A M O V IN G T AR GE T Ecosystem science has developed rapidly in the last 20 years, replacing old paradigms with new paradigms. We have learned, for instance, that succession seldom proceeds monotonically to a characteristic climax (Pickett and McDonnell 1989) and that productivity is not necessarily related to diversity (Waide et al. 1999). Our understanding of wetland function has undergone a similar, often parallel, development. In the early 1970s, when wetland science was still in a nascent stage, wetlands were ecotones between land and water, and freshwater wetlands were destined to fill up and become forests (Shelford 1963). Our slowly growing understanding of the control that the wetland environment has over vegetation species composition was hastened dramatically by the revelation that the soil environment becomes more reduced with longer periods of flooding (Ponnamperuma 1972), helping to explain many patterns of wetland vegetation zonation. We now know that many wetlands are ecotones not only between land and water-bodies, but also between surface ecosystems, both terrestrial and aquatic, and groundwater ecosystems (Novitzki 1978; Wright 1992). Like many terrestrial ecosystems, wetlands often have no well-defined successional sere (Niering 1989; Mitsch and Gosselink 2000). Disturbances such as periodic fire, wind or flood may instead maintain the characteristic structure and function of a wetland, with the result that many, if not most, will never become terrestrial ecosystems in the absence of climate change (Ewel 1990b; Janssens et al. 1992). Additionally, vegetation itself can affect redox relationships dramatically (Carlson and Yarbro 1988; Gleason et al. 2003), in some cases leading to tightly coupled feedback systems. The importance of disturbance to wetlands is now widely accepted. Flooding is a universal disturbance, but high winds, fire and the rise and fall of animal populations also affect many wetlands. Because of the importance of flooding in particular, the effect of a disturbance on a

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wetland is often mediated by characteristics of the catchment within which a wetland lies (see Acreman and McCartney, Chapter 28). Keeping up with emerging ecological principles that describe wetland structure and function, and the role of ecosystem-specific or region-specific disturbances, have been challenges for professional ecologists. Keeping the public apprised of new ways of thinking has been even more difficult, because critical components like groundwater are hidden from view, and important variables like redox chemistry are not easily measured and can be difficult to explain. The public’s expectations of how a wetland might serve humanity may sometimes be based on outdated paradigms. The most common cause of the failure of wetlands to provide expected services may therefore be unrealistic expectations. Policy makers, as well as stakeholders, may not be aware of advances in wetland ecology, remembering instead the older, simpler explanations of how familiar wetlands function. In this chapter, I emphasise the importance of clarifying essential relationships that may not be widely understood, and I introduce possible future changes in our understanding as the prospect of climate change – and continued change in wetland structure, function and services provided – becomes more certain.

ACCOUN T IN G FOR SPAT IAL AN D T EMPOR AL V AR IABILIT Y IN WET LAN DS Wetlands with the same descriptive name, such as bog, may vary considerably among themselves, and even the same wetland may have different characteristics from one year to the next depending on climate and other disturbances. The hydrogeomorphic wetland classification scheme outlined in Brinson (Chapter 22) distinguishes seven general types of wetland based on dominant water source, hydrodynamics and geomorphic setting, which help to group wetlands by common characteristics. None of the

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Introduction – How Do Wetlands Fail? wetland values listed in Table 27.1 is relevant to only one of these types. Some types of wetland span several geomorphic units. For instance, cypress (Taxodium distichum) swamps, which are forested wetlands common in the south-eastern USA, can be lake fringe swamps, floodplain forests, strands (slowly flowing channels that are sometimes dry), ponds and savannas (Ewel 1990b). These five types can fall into four hydrogeomorphic classes: lacustrine fringe, riverine, depressional and slope. They also differ somewhat in the services they provide to society. Several are sources of wood products, a few have been used for wastewater recycling, and many are habitat for different kinds of wildlife (Ewel 1990a). The value of even a single type, such as cypress ponds that can be used for wastewater recycling, may differ from one site to another depending on the size of the swamp, the size of the population proposing to use it, the distance of the swamp from the population, and the pattern of use of the swamp by wildlife. Nevertheless, to a lay stakeholder, a wetland can fail to provide a service even though its ability to do so has not been determined in the first place. Whether a particular type of freshwater wetland can provide a service such as wastewater treatment, or produce commodities such as commercial timber, furbearers (mammals exploitable for fur), waterfowl or a diversity of wildlife, in general depends a great deal on what kind of wetland it is and on its spatial and temporal variability. Avoiding surprises and subsequent failures means knowing how the wetland might respond through time to climate variables such as drought and wind as well as to fire and changes in animal densities. Variability caused by climate Services provided by a given wetland can be affected dramatically by both seasonal and longer-term changes in water, temperature and wind. Much of the early interest in using wetlands for improving water quality focused on the

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use of freshwater marshes for sewage disposal, but nutrient uptake in many cases was limited to only part of the growing season, usually spring and early summer (Simpson et al. 1983). Problems with temporary failures of other services can be predictable. Coastal wetlands are often valued for their ability to protect developments from storm tides but, during times of high water, coastal wetlands along the Great Lakes are breached and eroded by storm tides (Keough et al. 1999). By the same token, the tsunami that struck South Asia in December 2004 caused extreme damage in many places, but its effects may have been mitigated in others, such as the Republic of Maldives, because it struck during low tide (Richmond and Gibbons 2005). Many wetlands are protected because of their importance to wildlife, but wildlife numbers may also be very sensitive to the occasional extremes in weather that characterise their environment. Moreover, not all wildlife responds similarly to the same environmental conditions. Alligators (Alligator mississippiensis) and a variety of wading birds in south Florida, USA, for instance, depend for a successful breeding season on regular rise and fall of water levels in freshwater marshes, whereas snail kites (Rostrhamus sociabilis) in the same area depend on marshes that are flooded for long periods of time (Kushlan 1989; Takekawa and Beissinger 1989). Drought concentrates fish in deep holes easily accessed by tactile feeders like wood storks (Mycteria americana), which are then captured by alligators. The same drought would create difficult conditions for snail kites, which feed primarily on an aquatic snail (Pomacea paludosa) that thrives in deep water. The northern prairies of mid-western North America are dotted with prairie potholes, which are marshes that occupy shallow depressions. Over a period of 5–30 years, a given marsh may be dry with emergent vegetation, flooded with emergent, submersed and free-floating grasses and herbs, or an open pond, depending on the presence or absence of drought, insects and diseases, and the local density of muskrats (Weller and Spatcher 1965; van der Valk and Davis 1978).

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A major drought can prevent many species of waterfowl from nesting or raising a brood, because standing water is necessary for both protection and production of the invertebrates that are essential to a chick’s diet (Batt et al. 1989). High densities of many wildlife species require intact wetlands with characteristic hydrologic regimes, but periodic droughts can reduce wildlife numbers to very low levels. In all these cases, a species’ ability to survive occasional extremes in climate depends to some extent on the effects on vegetation, as well as the size of its own population before the onset of the perturbation. Some wetlands are subject to periodic high winds, most notably mangrove forests in the Caribbean, which experience hurricanes every 20–30 years (Lugo and Snedaker 1974). Freshwater, forested wetlands can be affected as well, but not as dramatically or predictably. Bay swamps (dominated by Persea borbonia, Magnolia virginiana, and other hardwood species) and cypress swamps in the south-eastern USA tolerate wind damage well, although other hardwood swamps do not (Putz and Sharitz 1991; Loope et al. 1994). Cabbage palm (Sabal minor), which is common in coastal groundwater swamps, survives storm tides, whereas southern red cedar (Juniperus silicicola) is easily uprooted and fractured (Williams et al. 1999). Many species in all these forests have proven commercially valuable at times throughout history. It is well known that steady winds reduce tree size and affect crown shape (Kozlowski et al. 1991), but there is no evidence other than the suggestion that frequent hurricanes may control mangrove forest structure (Lugo and Snedaker 1974) to indicate that exposure to regular episodic winds controls tree size in freshwater wetlands. Marshes may show less obvious effects from winds or associated storm tides, but the effects of wrack deposition by such storms could be substantial, long-lasting and highly variable (Keough et al. 1999). Response of vegetation and, subsequently, animal populations from any of these disturbances can therefore be distinctive and unpredictable. The ways in which weather affects wetlands are likely to change with the intensification of

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climate change. In addition to sea level rise and general warming, climate change may increase the frequency of tropical storms, or at least increase the impact of storm surges associated with these storms (Raper 1993). In some areas, winds that accompany severe weather are likely to become stronger, whereas rain events may become more intense, although total rainfall could increase or decrease in different areas (Hennessy et al. 1997). The effects of drought and wind are therefore likely to become even less predictable over the next several years. Climate change will probably affect northern wetlands more than wetlands at lower latitudes, and is likely to change the rate of methanogenesis (Bridgham et al. 1995). Failure of such globally important services as carbon sequestration by northern peatlands will have been caused by our own unrealistic expectation of how much atmospheric variability can be tolerated. Variability caused by fire Variation in fire frequency and in the pattern of recovery of vegetation from a fire may also be irregular and thereby affect the delivery of ecosystem services. Although much destruction of wetlands is associated with excessive burning, including peat fires, many wetlands depend on at least occasional burning to retain essential characteristics (see Joosten, Chapter 30). Periodic burning determines whether a shallow depressional wetland in the south-eastern United States produces commercially important pond cypress (Taxodium distichum var. nutans) trees or seldom-harvested bay trees such as Gordonia lasianthus (Casey and Ewel 2006). It keeps woody species from growing into tidal fresh marshes along the southern USA Gulf Coast and altering wildlife habitat (Nyman and Chabreck 1995), and it affects both species composition and palatability of vegetation to wildlife (Kushlan 1990). Cyclic changes in both fire and hydrologic regime can generate high levels of species richness in ecotones of wetlands, often the preferred habitat of rare species and therefore an important source of biodiversity (Kirkman 1995).

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Introduction – How Do Wetlands Fail? Controlled burning appears to preserve ecosystem services such as support of biodiversity, although the practice of burning to enhance rather than sustain wildlife habitat needs to be re-examined as changes occur in the importance of such services as production of game species (Nyman and Chabreck 1995). Nevertheless, finding desirable and sustainable fire management practices is especially important as natural patterns disappear with fragmentation of the landscape. Climate change will further affect fire frequency, particularly where droughts increase in frequency and stream flows and lake levels decrease. In boreal wet meadows, for instance, fires will probably increase in intensity, frequency and severity, which could lead to a shift in dominance from a native grass to a Eurasian species and, in turn, affect the productivity and diversity of large and small mammals (Hogenbirk and Wein 1995). Variability caused by animals The direct effects of animals on wetlands are particularly obvious when they affect drainage. The beaver (Castor canadensis and C. fiber) is notable for altering drainage networks in the USA and Europe by building dams that temporarily increase the area of ponds and riparian vegetation, leaving snags (dead trees) and other forms of landscape heterogeneity long after the dams have broken (Pollock et al. 1995). Movement of nutrients to downstream communities, affecting vegetation in both source and receiving ecosystems, has a long-term impact on the entire drainage basin (Naiman et al. 1994). Animals can affect other organisms without necessarily altering hydrology. Activities of the muskrat (Ondatra zibethicus) can combine with periodic droughts and floods to generate substantial variation in species composition and habitat characteristics in freshwater marshes in the mid-western USA (van der Valk and Davis 1978). The American alligator plays the same role in freshwater marshes such as the Everglades in

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Florida, USA, where the ponds it creates, described above, are often important feeding areas for fish and birds (Kushlan 1974). The connection between wetlands and wildlife production is important to many people, but sustaining it may require careful balance between wildlife population control and maintenance of appropriate patterns of hydrology and fire.

COMPEN SAT IN G FOR IN FLUEN CES FR OM T HE SUR R OUN DIN G LAN DSCAPE By their very nature, wetlands depend on both surface and groundwater inflows for their existence, but these in turn depend heavily on conditions in the catchment. Along with water comes a variety of organic and inorganic chemicals: some are essential nutrients, but others are pollutants (including essential nutrients in superabundance). Changes in water regime or nutrient concentrations alone can alter the basic nature of a wetland, as detailed particularly in Acreman and McCartney (Chapter 28). Invasions of both native and non-indigenous species may be promoted by changes in hydrology, nutrients or other pollutants (Finlayson, Chapter 29). Depending on a wetland for a specific service therefore means depending also on constancy in the pattern of inputs from the surrounding landscape. Hydrologic influences The results of changes in water regime are often profound, even though they are not always immediately obvious. In south Florida, USA, in the 1980s, for instance, surface water that was discharged from agricultural and urban lands into the Everglades resulted in changes in water levels that were unseasonable in both timing and magnitude. Breeding and feeding habits of many organisms depend on a predictable pattern of water level fluctuations (as described earlier for the American alligator, wood stork and snail kite). Slow declines in these and other animal populations are not easy to detect, and it is not

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always possible to reverse environmental conditions quickly enough to rescue a population or even a species. Extinction of the dusky seaside sparrow (Ammospiza maritime nigrescens) over a period of 60 years cannot be attributed with certainty to any single cause; mosquito control efforts, including impoundment and spraying for DDT, have been implicated (Montague and Wiegert 1990). Additional impacts of water regulation on vegetation and wildlife are described in Acreman and McCartney (Chapter 28). Decreased water inflow can be just as significant as increased water inflow. Along the Platte River system in Nebraska in mid-western USA, water diversion for irrigation and reservoirs caused wetland loss, stabilisation and narrowing of river channels, and led to establishment of woodlands dominated by Populus deltoides and Salix amygdaloides along with other species (Johnson 1994). These kinds of change have not benefited the wildlife that have used the area for generations, including rare species such as the whooping crane (Grus americana) that require wide, active channels (United States Fish and Wildlife Service 1981, cited in Johnson 1994). Small or gradual changes in hydrology can affect vegetation structure through changes in regeneration patterns. Because fewer species can tolerate increased hydroperiods, vegetation composition is likely to change more rapidly following decreased rather than increased water inflow, since a much greater variety of species can become established. Organic and inorganic pollution Changes in nutrient regime are usually due to enrichment (eutrophication) by nitrogen or phosphorus as a result of fertilisation in the catchment. Agricultural inputs from fields or feed lots and horticultural applications from lawns or gardens may be responsible. Wetlands, like lakes, often respond rapidly to nutrient enrichment with increased growth of vegetation (Schindler 1975; Nessel et al. 1982). Increases in productivity by algae, periphyton and herbaceous macrophytes have a dramatic effect on overall species diversity

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and food chain relationships (Turner et al. 1999). The ecosystem services of wildlife production and support can be affected dramatically by agricultural pollution, as demonstrated by the bioaccumulation of trace elements such as arsenic and selenium in wetlands that receive irrigation drainage in the western United States (Lemly 1994). Migratory waterfowl and other wildlife that are drawn to these few remaining wetlands subsequently suffer a variety of toxic effects. Invasive species Like most types of ecosystem, wetlands have been invaded by several species of aggressive plants within recent years (see Finlayson, Chapter 29, for specific examples). Many of these plants are non-indigenous, and their effects are accentuated by changes in environmental conditions and lack of herbivores. Novel phenotypes may appear under conditions that would not have been favourable to either parent (Galatowitsch et al. 1999). Other aggressive species are native but are favoured by hydrologic alterations. For instance, water level stabilisation in the Great Lakes has encouraged invasion of woody herbs and shrubs, both native and non-indigenous, into marshes normally dominated by grasses and herbs (Keough et al. 1999). These changes are particularly common in urban wetlands, which are subject to a variety of changes in their catchments (Magee et al. 1999). Invasive animal species can also have serious effects. Nutria (coypu, Myocastor coypus) were introduced for commercial fur production into freshwater marshes in Louisiana, USA, from South America. Heavy grazing by these large rodents immediately following pulses of muskrat herbivory can slow considerably the regeneration of vegetation (Chabreck 1988). Even when an invasive species can be eradicated, it may cause changes that continue to affect the kinds of ecosystem services provided by a wetland. Increased soil erosion, changes in drainage channels and subsequent salt water intrusion into wetlands was caused by grazing and trampling by the introduced Asian water buffalo

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Introduction – How Do Wetlands Fail? (Bubalus bubalis) in Kakadu National Park, northern Australia, a UNESCO World Heritage site and an important source of traditional foods for Aboriginals (Roberts 1995). Removal of the water buffalo led to an increase in plant biomass and to an apparent shift in the fire regime, caused both by changes in biomass accumulation patterns and by park managers (Roberts 1995). Understanding why a wetland does not deliver a particular service, such as support for a traditionally common wildlife species, may therefore require a look beyond the boundaries of the wetland as well as back through time. Changes in nutrient regime and other forms of pollution often accompany changes in hydrologic regime. Changes in fire regime caused by conversion of adjacent land to agricultural fields or by construction of fire breaks may have a substantial effect on a wetland, even though the land use change that generated it originated in the terrestrial part of the catchment. Accounting for spatial and temporal changes in a wetland throughout an entire landscape may therefore help explain why wetlands sometimes fail to meet our expectations.

RE ST O R IN G AN D CR E AT IN G W E T L AN DS Restoration of degraded wetlands and creation of wetlands de novo as mitigation for wetlands destroyed or altered in development have become major industries in the USA, as we seek to reproduce ecosystem services that have been lost. However, many of these mitigation wetlands have failed us too, and the outlook for both restoration and creation is not entirely promising. Success must be evaluated from the perspective of the target. When re-establishment of vegetation cover and standing water is desired, many projects have been successful. On the other hand, when recolonisation by organisms that depend on the soil to generate specific habitat characteristics or food items is necessary for success, complete restoration of a wetland may require decades or centuries. If decades must

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pass before success can be judged, at what point can failure be declared? Because of the stochastic nature of many hydrologic processes (and wetland responses), restoration even of marshes may take more than five years (Mitsch and Wilson 1996). Successful restoration after 11 years of a stand of trees characteristic of a riverine forest, averaging nearly 12 m in height, with basal area of 8.2 m2 ha−1, has been reported (Clewell 1999). However, on a larger scale, restoration of forested wetlands throughout the Lower Mississippi River alluvial valley of the USA remains problematic after 10–15 years (King and Keeland 1999). Failure to achieve consistent success is attributed to such factors as inappropriate hydrologic regimes and inadequate production of seedlings by nurseries. Complete restoration of characteristic wetland structure and function, including viable wildlife populations, clearly will take several decades. Landscape issues also restrict the degree of success of restored and created wetlands in providing services. In the USA, regulations sometimes permit wetlands to be destroyed in urban areas, where ecosystem services are more likely to be realised but where many developments are located, on the condition that they are replaced in rural areas where low population density increases the availability of land for wetland creation (King and Herbert 1997). However, few restoration or creation projects take into account the importance of having a characteristic landscape signature as a target (Bedford 1999; Houlahan et al. 2006). The nature of a wetland depends heavily on the landscape in which it sits for maintaining appropriate patterns of inflow and outflow of water and nutrients. Moreover, organisms within individual wetlands may require not only appropriate upland habitat to complete their life cycle, but additional wetlands as well. Restoration of a single wetland in a landscape, rather than the entire original complement, can reduce the likelihood of success, because organisms may not be able to disperse effectively in the altered landscape (McCauley and Jenkins 2005). Aiming for success in restoring or creating a wetland may be aiming at a moving target in many regions. Movement

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of animals into and out of wetlands affects other ecosystems too, and the failure to understand what role any but the largest and most obvious animals play suggests that success in being able to rescue these ecosystems with confidence is not yet within our reach.

D E PE N DIN G ON W E T L AN DS Most of the world’s wetland losses (56–65% of available wetlands) have been to drainage for intensive agriculture (Finlayson and Davidson 1999). Rice paddies that have replaced freshwater wetlands throughout Asia, in some places for millennia, continue to support millions of people. In the USA, rich soils that underlie drained wetlands in the mid-western part of the country have supported most of the country’s grain

production for over a century. The problem with stemming the tide of wetland conversion, even in today’s enlightened climate, is that for some countries development of a wetland still offers the hope of similar economic gain. Developed countries with high concentrations of wetlands, like Canada (where wetlands cover about 15% of the landscape), can promote ‘wetland industries’ in such economic sectors as ‘products and manufacturing’ and ‘supplies and distribution’ as well as ‘services’ and ‘knowledge’, without engendering the loss of additional wetlands (Warner 2003). At the other end of the scale, however, are developing countries such as the Federated States of Micronesia, where wetlands may cover as much as 30% of the total land area and support a substantial subsistence economy (Drew et al. 2005). Unfortunately, it is in countries such as these that policy makers

Box 27.1 Greentree reservoirs: an ecosystem failure? In the 1930s, wildlife managers in the southern USA began impounding wetlands dominated by deciduous trees, primarily oaks, to increase habitat for migrating waterfowl. By constructing a low levee around an area of bottomland forest and maintaining 30–40 cm of water from the time when trees become dormant in the fall to just before they flush again in the spring, acorn production was increased (Minckler and McDermott 1960) and growth rates of some of the trees were increased (Broadfoot 1967) for several years. These ‘greentree reservoirs’ proved to be very attractive to ducks during the hunting season, (Figure 27.1) and they appeared to mitigate the rapid loss of naturally flooded bottomland forest habitat (Guntenspergen et al. 1993). After several years, tree mortality proved to be very high (Wigley and Filer 1989), and shifts towards more flood-tolerant species of tree and other vegetation were apparent (King et al. 1998). It is clear that a longer hydroperiod, even during the dormant season, cannot be tolerated by these hardwood swamps. More

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careful management, such as by incorporating increased variability in the water level regime, may allow a novel type of wetland to be created and maintained as partial mitigation for loss of native bottomland hardwood stands (King and Allen 1996).

Fig. 27.1 A duck hunter in a greentree reservoir on the Noxubee National Wildlife Refuge in Mississippi. (Photo by D.J. Moorhead, University of Georgia, Bugwood.org.)

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Introduction – How Do Wetlands Fail? are more likely to risk sacrificing the basic and valuable goods and services offered by wetlands for an economic boost that will be able to support replacement goods and services as well as many more. Indeed, early reports from South Asia indicate that in some areas coastal redevelopment following the tsunami disaster does not include restoration of mangrove

Box 27.2

forests in spite of abundant evidence that these ecosystems would protect against future tsunamis and storm tides (Check 2005). If wetlands are not to fail society again, it is important that policy makers understand the full complement of goods and services offered by wetlands and the nature and magnitude of the risks involved in their loss.

Mercury in Wetlands: an ecosystem failure averted?

For more than a century, runoff from agricultural fields has been flowing into the Everglades of south Florida, USA, and for many decades the vastness of the Everglades system may have seemed adequate for ‘absorbing’ whatever pollution it might be receiving. This large and important wetland has, however, been subjected to more than eutrophication. Since the 1920s, water that might once have flowed into and through the Everglades has been redistributed by a system of canals, and discharged either into the Everglades or offshore, depending on management guidelines at the time. For decades levees constructed in the wettest parts of the north Everglades have impounded water for flood control and water storage in three water conservation areas. In the early 1900s, mercury contamination in fish and other vertebrates was detected in the Everglades, raising concern about the safety of any organism harvested for food in a region once considered to be remote from human influence (Figure 27.2). Although much remains to be determined about the details of mercury movement through wetlands in general, a clearer picture is beginning to form about the source of this problem in the Everglades. Mercury volatilises readily into the atmosphere, especially on hot summer days, and remains there for approximately a year before returning, primarily in rain, often far from its source. The sources of most mercury that enters the Everglades in atmospheric deposition are mining, smelting and the burning of

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fossil fuels and municipal and medical wastes (Atkeson and Axelrad 2003). Once mercury has been deposited, it has a very high affinity for sulphide, and sulphate-reducing bacteria may be responsible for the formation of methylmercury, particularly under conditions of low pH, moderate sulphate concentrations and high chloride concentrations; methylmercury is the most toxic and most readily bioaccumulated form of mercury (Morel et al. 1998). The fact that periphyton (algae and associated microorganisms that grow in mats attached to vegetation, rock and other surfaces, as in the photograph above) rather than phytoplankton, is the base of the Everglades food chain (Browder et al. 1994) may be important in explaining bioaccumulation of mercury in this vast wetland. High concentrations of methylmercury, characterised by considerable spatio-temporal variability, may result from complex interactions among several factors including seasonal changes in the types of fish present at a site, their diets, total mercury concentration, sulphate concentration and redox levels in surficial sediments (Cleckner et al. 1998; Gilmour et al. 1998). The development of toxic concentrations of mercury in Everglades fish and other vertebrates may therefore be related only indirectly to agricultural activity at the top of the Everglades catchment. Instead, increased atmospheric inputs of anthropogenic mercury from a variety of sources, many of them outside Florida, may be combining with nutrient and redox

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inter-relationships, which result as much from drainage and impoundment as from agricultural pollution, to affect an unusual food chain in a way that may be unparalleled elsewhere on earth. Increased regulation of mercury emissions in Florida since the early 1990s has led to substantial declines in the atmospheric deposition of mercury (Atkeson and Axelrad 2003). Concentrations of mercury in the feathers of wading birds have also declined (Frederick et al. 2002), and the threat of mercury exposure to the Florida panther, an endangered species that often feeds on piscivorous wildlife species such as raccoons, is also substantially lower (Barron et al. 2004).

Box 27.3

A developing Country on a Wetland landscape: an ecosystem failure waiting to happen?

The Federated States of Micronesia (FSM) is a nation whose boundaries encompass more than 2.5 million km2 in the western Pacific Ocean, but its more than 600 islands cover only 700 km2 of land. Once part of the Trust Territory administered by the USA after the Second World War FSM has been independent since 1979 and is now linked to the USA by a Compact of Free Association. Kosrae, the easternmost of the four FSM states, is a small (109 km2) island characterised by high rainfall (>5 m a−1) and a rugged, mountainous interior. Approximately 30% of its land area is wetland, including both mangrove forests and freshwater forested wetlands, as well as a few marshes and montane bogs. In spite of the funding available to FSM from the Compact, the 7700 or so Kosraeans depend heavily on a subsistence economy, harvesting food and timber from wetlands at a rate amounting to approximately 60% of the median household income (Naylor and Drew 1998; Drew et al. 2005). Most of these products come from a simple wetlandbased agroforestry system (Figure 27.3) that has apparently been in place for more than a millennium (Athens et al. 1996) and still appears to be sustainable (Chimner and Ewel 2004). Unfortunately, a subsistence economy cannot

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Fig. 27.2 The Everglades ‘river of grass’ hides quantities of mercury, which can bioaccumulate in higher organisms. (Photo by K.C. Ewel.)

support schools and hospitals or purchase fossil fuel to produce electricity. Participating in the global economy means producing exports or services that can bring in external funding, and in a wetland-dominated landscape, wetlands are usually at risk in any proposed development. As the demands of participating in the global economy grow and grow, the earth’s remaining wetlands, especially in the tropics where so many developing countries and so much wetland area are located, are increasingly threatened.

Fig. 27.3 Taro growing in a simple wetland-based agroforestry system. (Photo by W.M. Drew.)

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Introduction – How Do Wetlands Fail? CO N CL U S IO N In too many cases, our sense of what a wetland should be depends on observations made over too short a period of time and across too narrow a landscape perspective. Wetlands will continue to fail us until we can understand more thoroughly what a service requires of that wetland, how variable a wetland is in its essential characteristics and in its ability to deliver that service, the role that the landscape around the wetland plays in sustaining these characteristics, and the economic importance of a wetland’s goods and services. Wetland conservation, restoration and creation will be better served when the ways in which different kinds of wetlands can benefit the many kinds of human societies are better understood. Nor can wetland conservation efforts be constrained by national boundaries. Developed countries must somehow ensure that no other country should have to sacrifice its own wetlands, or other important ecosystems, to meet the demands of the global economy which is, in turn, controlled by developed countries.

ACK N OW L E D G E M E N T S I thank Julie S. Denslow and C. Max Finlayson for reviewing the manuscript.

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28 Hydrological Impacts in and around Wetlands M I C H A E L C . A C R E MA N 1 AND MATTH EW P. McCARTNEY 2 1Hydro-Ecology

and Wetlands, Centre for Ecology and Hydrology, Oxfordshire, UK Water Management Institute, Addis Ababa, Ethiopia

2International

IN T R O D U CT ION Wetlands are dynamic systems. They continually experience change as a consequence of natural phenomena, for example deposition of sediment and organic material, subsidence, drought and, in the case of coastal wetlands, sea-level rise. Wetlands are transient features of the landscape. They change over time and eventually disappear, whilst new wetlands are created elsewhere. The natural process is for them to in-fill in the long term with sediment and vegetation and to become gradually drier. Successive colonisation by floating, emergent and fringing plants and, finally, terrestrial plants may complete the succession from aquatic to land-based ecosystems. Their functions, therefore, may also change over time. Site management of wetlands may consequently be required to maintain the level of desirable functions rather than to allow natural succession. At the same time, other wetlands will be forming naturally, as a result of newly formed meander cutoffs, shifts in coastal sand bars or silting of river channels. In the past, the effect of human activities on wetlands was generally insignificant and of a local nature. In many cases, the wetlands had sufficient resilience to recover from the human induced stresses placed upon them. However, the past 200 The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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years has seen an accelerated and unprecedented loss of natural wetlands due to direct and indirect human activity. Increasing population, coupled with technological advances and intensive agricultural development, have had an ever-greater impact on wetlands. Changes in water balance and water quality have resulted in environmental degradation, destruction of natural habitat and loss of ecological functions, with serious implications for the integrity of many wetlands and the environmental benefits that they provide. To some extent, we have created new, artificial wetlands by building reservoirs, canals and flood storage areas. These may go some way towards compensating for the loss of natural wetlands. For example, in the United Kingdom, reservoirs provide an alternative habitat for wildfowl and other water-associated organisms, now that extensive areas of natural wetland have been drained. The general importance of these reservoirs to wildlife conservation is indicated by the fact that, out of a total of 500 water supply reservoirs larger than 20 ha in England and Wales, 174 have been designated as Sites of Special Scientific Interest (Moore and Driver 1989), including Rutland Water, which is also a Ramsar site (a Wetland of International Importance under the terms of the Convention on Wetlands). Artificial fishponds are a key component of the Hortobágy Ramsar site in Hungary. However, in general, artificial wetlands support less biodiversity, and overall wetland losses and the less tangible effects of degradation of functioning have far outstripped the gains.

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It is the presence of water for some significant period of time that creates the characteristic features of a wetland. Hydrological processes are closely linked to the chemical and physical aspects of wetlands, which in turn affect biotic functions (see Baker et al., Chapter 6; Gilvear and Bradley, Chapter 7). Hydrological processes directly influence species composition and richness, primary production, organic matter decomposition and the cycling of nutrients in wetlands. Consequently, the modification of hydrological patterns through human interventions can result in large-scale changes in the ecological status and functioning of wetland ecosystems. Humaninduced changes to the hydrological regime are amongst the most significant impacts that wetlands experience (Millennium Ecosystem Assessment 2005). Until recently, in many places wetlands were widely perceived as wasteland that provided breeding grounds for mosquitoes. In much of Europe and the USA draining of wetlands was viewed as a benefit to society and encouraged by laws and subsidies. Worldwide, conversion or drainage for agricultural development has been the principal cause of inland wetland loss (Millennium Ecosystem Assessment 2005). The amount of wetland lost is difficult to quantify because detailed inventories generally do not exist and definitions of ‘loss’ are subject to a wide range of interpretations, however, some estimates of wetland loss have been attempted. Many European countries have lost between 50% and 70% of their wetlands in the last century (McCartney et al. 2000a). The United States has lost some 54% of its original 87 million hectares of wetlands (Tiner 1984), primarily to drainage for agricultural production. In Asia some 67% and in Latin America and the Caribbean 50% of the major threats to wetlands were from hydrological change related to drainage for agriculture, pollution, catchment degradation or diversion of water (WCMC 1992). Alterations to any part of the hydrological cycle can have significant impacts on wetlands. Such alterations can take many forms, ranging from changes to climatic inputs, changes to water movement through the soils and rocks of a catchment or direct control of water movement in and

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around the wetland. Hydrological impacts in and around wetlands can be divided into six areas: • variations in climate, including precipitation and temperature (either natural or human-induced), • changes to the catchment upstream of the wetland (e.g. urbanisation, deforestation, land drainage), • river engineering (construction of dams and embankments that alter river flows), • pollutant discharges (from industry, agriculture and domestic waste), • groundwater abstraction (that reduces the water table), • direct wetland management (drainage, mineral extraction, pumping into or out of wetlands, or changes in wetland vegetation).

CLIMAT E CHAN GE Recent drought events and the clustering of particularly warm months and seasons have highlighted the concern that the climate is changing significantly and will change even more over the next century. The Intergovernmental Panel on Climate Change (IPCC 2001a,b, 2007) has concluded that the balance of evidence suggests a discernible human influence on global climate and therefore that the global warming of the last 100 years is certainly not entirely due to natural causes. The HadCM2 model (run by the UK Hadley Centre) predicts that, with unmitigated emissions of greenhouse gases (the ‘business as usual’ scenario), global temperatures will rise by 3°C by 2080 compared to the present (DETR 1999). High latitude areas and northern South America, India and southern Africa are expected to warm more quickly than the global average, increasing wetland evaporation. A rise in sea level of 40 cm is expected by the 2080s, producing major impacts on coastal wetlands. Reduced rainfall of up to 50% is predicted for many areas including central America and northern South America and Australia, and over 50% in July in north Africa and southern Africa (Figure 28.1). This will result in reduction in rainfall, river flow and groundwater inputs to wetlands by up to 50% in the worst hit areas. For the Mediterranean region, Estrela et al. (1996) have estimated that a 1.5–2.0°C rise

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< –50 –50–0 0–50 50–100 100–200 200–300 > 400

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Fig. 28.1 Changes to annual precipitation in 2080 compared with 2000 as predicted by the Hadley Centre CM2 model with ‘business as usual’ greenhouse gas emissions scenario.

in temperature could result in a 10% reduction in rainfall and a 40–70% reduction in renewable water (river flows or groundwater recharge). In contrast, more rainfall is predicted for central Africa, China and North America. There have been few studies that have quantified the impacts of climate change on wetlands, although some publications have highlighted the threat, such as in India (Foote et al. 1996) and China (Shen and Cao 1994). Studies of climate change on semi-permanent prairie wetlands in North Dakota (Poiani et al. 1995) showed that they were most sensitive to changes in spring precipitation. Burkett and Kusler (2000) recognized that not only is climate change likely to

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lead to loss of wetlands, such as tundra, marshes and wet meadows underlain by permafrost, but wetlands that are dried can become net sources of carbon dioxide (but with possible reduction of methane) serving as a positive feedback to global warming. Parish et al. (2007) came to similar conclusions for peatlands. In Great Britain, Acreman et al. (2009a) suggest that predictions of reduced summer rainfall and increased summer evaporation will put stress on wetland plant communities in late summer and autumn with greater impacts in the south and east of GB. In addition, impacts on rain-fed wetlands will be greater than on those dominated by river inflows. In the UK, a hydrological feasibility study has been undertaken

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to restore 37 km2 of former fenland (Mountford et al. 2002). Under current winter pumping conditions, there is an annual water deficit of around 3 million m3. Modelling of climate change scenarios for the 2080s (using the UKCIP medium-high forecast, which includes a reduction in summer rain of 39%, an increase in summer evaporation by 48% and an increase in winter rainfall of 21%) show that this deficit could increase to 10 million m3. Coastal wetlands in particular will be at threat from sea level rise, such as those in the Alligator River region of Australia (Finlayson et al. 1997) and the Mediterranean area of Europe (Nicholls and Hoozemens 1996). Any degradation of the wetland due to climate change may have implications for its ability to perform important functions. For example, Freeman et al. (1993) showed that a lowered water table anticipated as a consequence of climate change reduced the efficiency with which valley-bottom wetlands acted as sinks or sources of nutrients. Climate change may exacerbate the impacts of other drivers impacting wetlands. For example, changes in climate may alter human water demand. Consequently, the impact of such changes on river flows and wetlands is a complex integration of both climate and other environmental and socioeconomic changes. Relatively small variations in temperature and rainfall can have significant impacts on evaporation and groundwater recharge. These changes will affect both total annual flows in rivers and their distribution through the year. In a dry area of Tanzania, model results indicate that a 15% reduction in rainfall would cause a 40% decrease in groundwater recharge (Sandstöm 1998), which will have serious implications for those rivers and wetlands in the region fed by groundwater.

C AT CHM E N T CHAN GE S The water supply to wetlands is derived partly by direct precipitation and partly as inflows from streams or groundwater. Both stream flow and groundwater levels will be affected by changes to the catchment upstream, including vegetation changes, urbanisation and agricultural drainage.

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Worldwide human-induced changes in landcover represent a very significant direct threat to wetlands. Changes in land-cover cause changes in the energy and material fluxes that support wetland ecosystems. It has been shown that conversion of forest to agriculture may alter the radiation balance of the surface, soil structure, evaporation and runoff generation (e.g. Gash et al. 1996). In the Amazon rainforest, 50% of rainfall is derived from local evaporation. After forest cover is removed, an area can become hotter and drier because water is no longer cycled between plants and the atmosphere. This can lead to a positive feedback cycle of desertification, with less precipitation and evaporation and an increase in surface temperature (Lean and Warrilow 1989) due to changes in albedo and roughness. Models suggest that rainfall is reduced by 26% for the year as a whole (Shukla et al. 1990). Similarly modelling of the removal of natural vegetation in the Sahelian region of Africa suggests that rainfall has been reduced by 22% between June and August and the rainy season has been delayed by half a month (Xue and Shukla 1993). In this region, wetlands, such as the inner delta of the Niger River, are particularly important areas of biological production providing livelihoods for local people (Zwarts et al. 2005). However, the impact on an individual wetland depends on the specific nature of interventions, the scale of the land cover change, and the interplay of site-specific factors such as soil type, geology and slope with local climate. It is widely recognised that vegetation promotes infiltration of water into the soil, aiding the recharge of underground aquifers, lowering flood risk and anchoring the soil, thus reducing erosion. Consequently, deforestation or overgrazing of grasslands has been shown to reduce infiltration and promote runoff, increasing flood risk and sediment load (Bruijnzeel 1990; Critchley and Bruijnzeel 1996). Such evidence has been used to suggest that deforestation in the Himalayas has caused increased flooding downstream on the floodplain and deltaic wetlands of the Rivers Ganges and Brahmaputra in Bangladesh (Agarwal and Chak 1991). However, whilst this impact can be demonstrated for small catchments and plot

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Hydrological Impacts in and around Wetlands experiments, there is little real evidence that it occurs on large catchments, since the impact is swamped by other factors (FAO/CIFOR, 2005). Kaimowitz (2001) claims that deforestation impacts on catchment hydrology are over-stated and that, for example, flooding resulting from Hurricane Mitch in Central America in 1998 was not exacerbated by deforestation. Much of the catchment of the River Ouse in eastern England was formerly fenland with seasonally inundated floodplains and permanent areas of open water reeds and raised mires (Figure 28.2). Drainage of the catchment, initiated in 1637 for intensive agriculture, has led to a lowering of the land surface by several metres through loss of peat soils. The remaining wetlands have been left elevated above the surrounding landscape and very vulnerable to drainage and groundwater pumping in the adjacent farmland. Wicken Fen, the first

national nature reserve established in 1899, is some 2 m above the surrounding farmland. The Fen is of high conservation value because of its unique and rich biodiversity and has been designated a Ramsar site. Historically, it was flooded each winter from the River Lode, which borders the site (Lock et al. 1997). In an attempt to stop drainage from the Fen into the surrounding low-lying farmland, an impermeable polythene membrane was installed along the perimeter of the Fen in the 1980s. The entire northern boundary of the Fen was waterproofed by the Anglian Water Authority (AWA) between 1987 and 1989 as part of a scheme for reducing water loss from the Lode system as a whole. It was recognised that water was moving from the Lode into the Fen and then, because of the higher elevation of the Fen, draining into the surrounding farms. The northern boundary of the Fen was therefore

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Evapotranspiration (mm)

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Interfluve forested (%) Fig. 28.3 Relative dry season evapotranspiration losses from the dambo and interfluve components, for different percentages of forest cover on the interfluve, in a catchment comprising 44% dambo and 56% interfluve. (After Bullock and McCartney 1996.)

adopted as the effective north bank of the Lode. To reduce drainage from the Fen, the ditches along Howes Bank and Spinney Bank were re-excavated and the spoil used to build up the banks. A vertical, 4 m wall of polythene was then inserted into the middle of the banks. This was done by digging out a narrow trench with an excavator, unrolling the sheeting and backfilling (Lock et al. 1997). As a mitigation measure, this has been reasonably successful, but partly because of its relatively small size, the Fen remains very vulnerable to changes in rainfall, river flows and groundwater levels. Only a few hydrological studies have been carried out on the impacts of planting broad-leaved woodlands, but it appears that in drought years forest evaporation is higher than from grassland or arable land, with resultant lower stream flow and reduced recharge to aquifers (Harding et al. 1992). At Thetford in UK, the forest cover reduced recharge by 50% (Calder 1992). It has been shown that the nature of vegetation surrounding seasonal headwater wetlands (dambos) in Zimbabwe significantly affects the relative proportion of evaporation from the wetland and the surrounding area (Figure 28.3) and so could affect catchment water balances (Bullock and McCartney 1996). Studies in India suggest that eucalyptus trees use about twice the amount of water as crops, but no more water than indigenous forests. Australian

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research found that when groundwater levels are high eucalyptus trees use no more water than other vegetation, but when the water levels fall their long roots enable them to continue tapping the groundwater, thus during droughts their water use is greater than crops’ (Calder et al. 1997). It has been proposed that deforestation of miombo woodland in southern Africa could increase dry season water resources, although it is recognised that this would have serious environmental consequences (Hough 1986). Evidence from Malawi indicates that miombo deforestation raises groundwater levels but, because of the resulting increase in upward discharge of groundwater, may promote gullying within wetlands (McFarlane and Whitlow 1990). Furthermore, although deforestation of small headwater catchments containing wetlands in Zambia increased dry season base flow, it also resulted in more rapid response and higher flood peaks in storm hydrographs (Mumeka 1986). What these studies indicate is that, at least at the small-scale, human modification of vegetation within a catchment can alter significantly the hydrological processes and water balance of wetlands. Improved understanding of the hydrological processes in a catchment has important implications for land use. Water management has become a key objective in land management. For example, South Africa has recently embarked on

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Hydrological Impacts in and around Wetlands a programme of removing non-native species, such as eucalyptus, to increase water resources, as these species are considered to deplete valuable water. It is estimated that alien plants cause the loss of almost 7% of the annual flow in South Africa’s rivers (Versfeld et al. 1998). In Australia, the opposite problem has occurred. Loss of eucalyptus trees has led to increased groundwater levels which have caused soil waterlogging and salinisation, so replanting is being encouraged. In Honduras, the La Tigra National Park, a 7500 ha cloud forest, sustains a high quality, wellregulated water flow throughout the year, yielding over 40% of the water supply of Tegucigalpa, the capital city (Acreman and Lahmann 1995). Because of its value for watershed protection, La Tigra has been the focus of an investment programme involving a series of economic incentives for villagers living in the buffer zones. The value of these services is considerable. In highly developed countries, such as the USA, ecosystem management is also being pursued to improve water resources. For example, rather than build water treatment facilities at a cost of US$7 billion, New York City is spending a tenth of this sum to ensure the protection of the biological and hydrological processes of the highlands of the catchment (Abramovitz 1997). Urban developments (with their large areas of roofs, roads, car parks and other impermeable surfaces) make runoff more ‘flashy’, thus flooding of wetlands downstream may be more frequent but of shorter duration. The same impact of faster runoff may result in catchments where soils are frozen or have been compacted by cattle. In Tanzania, pastoralists have been excluded from the Usangu wetland because, although not proven scientifically, there is widespread belief that over-grazing was degrading the wetland and reducing its ability to hold and release water (Lankford et al. 2004).

R IV E R E N G IN E E R IN G River engineering projects can have very significant impacts on the hydrological regime and hence

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on wetlands downstream, both freshwater and coastal. Dams in particular are designed to store water during the wet season for subsequent use during dry periods when required for hydropower generation, irrigation or domestic and industrial supply. Not only are upstream river ecosystems lost because of inundation, but the downstream river flow pattern is altered. Over half (172 out of 292) of large river systems are affected by dams, including the eight most biogeographically diverse. (Nilsson et al. 2005). Downstream this may lead to changes in water temperature and chemistry, sedimentation and channel morphology. Thus, dams modify, in very complex ways, the conditions to which local ecosystems, including wetlands, have adapted. These changes can have very great impacts on the natural biota, sometimes many hundreds of kilometres downstream (McCartney et al. 2000b). The construction of the Kariba Dam on the Zambezi River has had a significant impact on the ecology of the internationally famous Mana Pools Game Reserve, located 130 km below the dam (Petts 1996). Under natural conditions, the active migration of the Zambezi River has produced a broad floodplain containing a series of residual pools. Prior to construction of the dam, the floodplain was inundated in most years to a depth of up to 5 m. Flooding persisted for 3 months or more; silts and nutrients were supplied and stagnant pools flushed. At the same time, large herbivores were driven off so that, aided by water-borne seed dispersal, the vegetation could recover from the intense grazing pressure to which it was subjected during the dry season. In this way, the productivity of the floodplain was replenished each year. Now, river regulation has reduced the ecological dynamism of the floodplain and removed the instability that was previously a key factor in the ecosystem, resulting in reduced productivity. Flood control during the wet season has induced a response in vegetation composition to one favouring drier conditions; the species composition of the grass sward has changed, with an increase in unpalatable herbs and grasses. Floodplain pools have become dominated with emergent rooted aquatics and water ferns. The elimination of

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flooding has allowed grazing to continue for longer periods than formerly, and some herbivores (e.g. water buck) have moved from their wet season dispersal areas to spend virtually all year on the floodplain. Consequently, the floodplain recovery period is markedly reduced and the habitat placed under additional stress. Overgrazing has resulted in habitat reduction and is believed to be the cause of reduced populations of hippopotamus, crocodile and various waterfowl. Acreman et al. (2000) highlighted the impacts of reduced inundation of floodplains and deltas caused by dams and the subsequent loss of wetland functions and products on which many rural communities depend, particularly in developing countries. The release of managed flood from dams to inundate downstream wetlands was recommended by the World Commission on Dams (2000) and was subsequently adopted as best practice for dam operation by the World Bank (Acreman 2002). In some cases, this loss of wetland functions exceeds the value gained by using the reservoir water for other purposes, such as irrigation. In north-east Nigeria, where the Hadejia and Jama’are Rivers combine within the KomoduguYobe basin, an extensive floodplain of around 2000 km2 used to be inundated annually. Since 1971, a series of dams have been constructed on the main tributaries including Tiga and Challawa Gorge. During recent droughts the area of land inundated had reduced, with only 300 km2 flooded in 1984 (Hollis et al. 1993). Water from the dams is used primarily to provide water for cereal irrigation. Hydrological modelling studies of the upper Niger River basin in west Africa showed that construction of the proposed dam at Fomi in Guinea could reduce water levels in the inner Niger delta during a flood by up to 60–70 cm (Zwarts et al. 2005). The most productive parts of the delta, in terms of fisheries, agriculture and pastureland, are shallow depressions on the floodplain fed by channels that connect them to the main river; reduced flood levels may mean that water does not reach the depressions (Acreman 2006). Barbier et al. (1991) undertook an economic analysis of a major irrigation scheme in the

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headwaters of the Hadejia River. They expressed their results in terms of value per unit of water, as this is the limiting resource in northern Nigeria, rather than in terms of land area, which had been used traditionally. They showed that the net economic benefits of the floodplain were at least US$ 32 per 1000 m3 of water (at 1989 exchange rates), whereas the returns from the crops grown on the Kano River project were only US$ 0.15 per 1000 m3 and when the operational costs were included, this dropped to only US$ 0.0026 per 1000 m3. Furthermore, this analysis did not include the other benefits of flooding, such as groundwater recharge or flows downstream to Lake Chad. At a meeting in Kuru in 1993, representatives from the responsible authorities including state water boards, River Basin Development Authorities, members of research institutes and government departments met to discuss the water resources of the Komodugu-Yobe basin. They agreed unanimously that ‘flooding in the wetlands made possible by managing releases from dams in the wet season should be maintained to make possible the production of rice, dry season agriculture, fuelwood, timber, fish, wildlife, as well as biodiversity and groundwater recharge’ (HNWCP/NIPSS 1993). In addition it was agreed that existing facilities at Tiga dam that could be used for artificial flood releases should be tested, the adequacy of outlets for releases from Challowa Gorge (dam) should be assessed and detailed reassessment of Kafin Zaki dam (being planned) should incorporate the capacity for adequate flood releases. Planners and decision-makers need to take account of wetland functions and values. Most decisions about development are made on economic grounds. However, traditional cost benefit does not consider ethical, political, social, historical or ecological issues (such as biodiversity), which cannot readily be given a monetary value. Multi-criteria analysis (MCA) provides a framework within which decisions can be made based on many measures, not just economic value. In this method, a range of criteria, including social and ethical considerations, may be used to support decision-making. Multi-criteria analysis has been used in the Po delta by Munda and Nijkamp (1995) and is recommended in

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Hydrological Impacts in and around Wetlands guidelines prepared by the World Commission on Dams (2000) to help make decisions between different development options. Other types of river engineering, including embankments and channel deepening, can have significant impacts on wetlands. Embankments separate the river from its floodplain, to protect agricultural land or urban developments. This prevents inundation of the floodplain during high flows, and the deposition of rich silt and exchange of nutrients between the river and floodplain. Many fish species such as pike (Esox lucius) in Europe (Wheeler 1998) and Tilapia in Africa (Welcomme 1976) breed on the floodplain, and reduced fish catch is directly related to reduced inundation (Figure 28.4). In many countries fish provide the main protein source for rural poor, and reduced floodplain inundation can significantly affect their livelihoods. A knock-on effect of preventing floodplain inundation is that floodwater storage is reduced and, thus, flood risk downstream may be increased. Modelling studies showed that embanking the River Cherwell in the UK (separating the river from its floodplain) increased flood peak flows by up to 150% (Acreman et al. 2003a). The magnitude of the 1994 floods on the Rhine was blamed partly on the loss of floodplain storage upstream. As part of the flood alleviation strategy, embankments were removed (Schropp and Jans 2000). Increasing the river channel capacity, by dredging, can have

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the same impact as embankments, reducing the frequency and extent of floodplain inundation, with similar implications. Some engineering schemes plan to divert water around a wetland to avoid losses of water resources by evaporation so that more is available downstream for irrigation or hydropower generation, such as on the Nile. Other schemes have been implemented for political reasons. For example, Maltby (1994) reports that the marshlands of Mesopotamia, in Iraq, have been seriously degraded, in terms of biodiversity and natural resources, by diversions of the Rivers Tigris and Euphrates. Nicholson (1993) suggests that this has had a significant negative impact on the Marsh Arabs who live there. The Sudd is the largest wetland in Africa and lies on the White Nile in southern Sudan dominated by Cyperus papyrus, Phragmities spp. and Typha spp. It was observed early in the last century that the outflow from the Sudd was only half the inflow, due to high evaporation rates (Sutcliffe and Parks 1999). Consequently, it is considered by downstream countries (primarily Egypt) to be the cause of losses of valuable water resources. In contrast, local communities regard the wetland as an important natural resource, particularly for grazing, and as a habitat for wildlife. The reduction in flow of the Nile led to proposals in 1904 to divert the river around the swamp through a canal to increase the supply of water downstream. Work has started several times on the Jonglei canal but stopped following internal warfare. Sutcliffe and Parks (1996) estimated that a constant discharge down the canal of 230 m3 s−1 (using the historical period 1905–1980) would have reduced the area of permanent swamp by 35% and the important seasonally flooded area by 23%. Analysis in two sample areas showed that the duration, maximum depth and range of inundation determine the vegetation distribution, which varies between deep-flooded species (Echinochloa stagnina, Vossia spp. and Cyperus papyrus) and the shallow-flooded species (Echinochloa pyramidalis, Oryza spp. and Phragmites). The proportions of different species over the floodplain are important because the local economy relies on grazing during the dry season of the two

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Echinochloa species and Oryza spp. At one end of the scale, the papyrus provides no grazing but is an important refuge for wildlife, while the Phragmites, which is limited to the dry end of the scale, is unpalatable. Between two surveys in 1951 and 1982, the rise of Lake Victoria caused river flows to double. This greatly increased flooding, and thus led to vegetation changes. Understanding of the mode of control makes it possible to predict the effects of such measures on the vegetation, which is vital to the local economy. It also suggests that lake and river management might be amended to minimise the adverse local effects (Sutcliffe and Parks 2001).

G RO UN DW AT E R AB S T R ACT ION Aquifers provide valuable water resources for people throughout the world. If, over a long period, more water is abstracted than is recharged to the aquifer, groundwater levels will fall and eventually the resource will be exhausted. In many areas, such as Jordan and Saudi Arabia, recharge is currently insignificant and groundwater that is abstracted is not being replaced. Aquifers here were charged some 10 000–20 000 years ago (during glacial phases in high latitudes) when rainfall was much higher and percolation of water to underlying rocks led to the build up of substantial groundwater resources (Goudie 1977). In many areas of the world groundwater levels are falling rapidly due to abstraction. Water tables are falling as much as a metre per year in many parts of China, India, Mexico, Yemen and elsewhere. In the Messara valley in Greece, groundwater levels have fallen by 30 m in 15 years, following installation of pumps to irrigate grape vines and olive trees, leading to complete loss of wetlands in the valley (Acreman 2000a). Even sites of international conservation significance are being degraded as a result of groundwater abstraction. Jordan is one of the world’s most water scarce countries, with rainfall of less than 200 mm per year in most areas. Consequently, underground water resources are crucial. The Azraq Oasis, designated as a wetland of international importance

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under the Ramsar Convention in 1977, is fed by the Azraq basin where rainfall is 350 mm per year. The Oasis has a permanent population of about 4000, who rely on seasonal salt extraction, irrigated farming and tourism. In addition, it is a traditional stopping point for caravans and around 40 000 nomads depend on it for water and grazing. Jordan also has a growing urban population. In 1991 55% of the people lived in or near Amman with a growth rate of 3.4%. The city depends on water from the aquifer underlying Azraq. By 1992 water withdrawals were 50 million m3 per year (a−1), twice the ‘safe yield’ of 22–24 million m3 a−1, and the water table fell from 2.5 to 6 m. The discharge rate of the four main springs fell rapidly from 10.5 million m3 a−1 to less than 1 million m3 a−1. As a result, in 1993 and 1994, the spring pools almost dried up and associated marshes diminished. Bushfires occurred more often, the salinity of the groundwater increased to 3000 parts per million, tourism and agriculture declined and out-migration rose. In 1994 the Azraq Oasis Conservation Project was established (Fariz and Hataugh-Bouran 1998) funded by the UNDP. A proposal was put forward to reduce groundwater pumping but this would have cut Amman’s water supply. Instead, the Ministry of Water and Irrigation agreed to pump some 1.5–2 million m3 a−1 abstracted water back into lakes in the centre of the oasis. This coincided with a wetter than average 1994–1995 wet season (with over 10 million m3 reaching the wetlands). This slowed salinisation, restored the wetland ecosystem and revitalised tourism. It also spawned a local community group, the Friends of Azraq Society, which promotes conservation and sustainable development. Unfortunately, the following wet season (1995–1996) was actually a severe drought with less than 30 mm of rain recorded. Despite the actions of the project, the water table has continued to decline and is more vulnerable to low rainfall episodes. This has further desiccated the wetlands and created new problems including a major fire hazard. It can be very difficult to predict the impacts of groundwater abstraction on wetlands. The interactions between a wetland and underlying

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Hydrological Impacts in and around Wetlands aquifers is normally unique in some way, making it difficult to extrapolate results of scientific studies from one site to another. For example, the Breckland Meres (Langmere, Ringmere and Fenmere) in the UK are superficially similar endorheic base-rich wetlands. However, their connection to the groundwater is quite different (Acreman and Jose 2000). Langmere is in hydrological continuity with a chalk aquifer and its hydrological regime is controlled by groundwater fluctuations (Figure 28.5). Ringmere has a less well-developed connection with the chalk due to a lining of organic matter, but is still largely controlled by groundwater. In contrast, Fenmere is separated from the chalk by clay alluvium and its water level is largely a reflection of the balance between rainfall and evaporation. This clearly presents a challenge for scientists to produce simple methods for assessing the impacts of groundwater abstraction on wetlands and indicates the need to monitor any wetlands where impacts are suspected. Guidelines produced for the International Convention on Wetlands (Acreman 2005) emphasise the need to develop a clear understanding of how a wetland works hydrologically, particularly where impact assessment studies are undertaken Water Sand Aquifer (e.g. chalk) Low permeability strata Organic layer

34

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Fig. 28.5 Breckland meres geology. (After Acreman and José 2000.)

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to assess the impact of groundwater abstraction on wetlands (Acreman and Miller 2007). For some wetlands that are linked hydrologically to an underlying aquifer, their functioning may vary depending on the prevailing conditions. When the water table is high the aquifer may supply water to the wetland (groundwater discharge), but as the groundwater level drops, the hydraulic gradient reverses and then the wetland may supply water to the aquifer (groundwater recharge). This is exemplified by the Tablas de Damiel wetland. The history of the Tablas de Damiel wetland in Spain provides an example of how groundwater exploitation has affected the interaction between surface water and the aquifer. The Tablas is a marshland at the confluence of the Rivers Guadiana and Gigüela. At its most extensive, it covered some 15 km2 with a depth of around 1 m. It is one of the two Spanish wetland National Parks. This status provides legal protection for the wetland itself but not for the catchment of the upper Guadiana which feeds it. The Tablas de Damiel has been designated a Ramsar site and nominated as a UNESCO Biosphere Reserve. The wetland is sustained predominantly by discharge from the western Mancha Occidental calcareous aquifer, although surface flow from the Guadiana and Gigüela also helps support the wetland. The aquifer has been intensively exploited for the past three decades, abstraction increased from 200 million m3 a−1 in 1974 to 600 million m3 a−1 in 1987. This latter figure is greater than the estimated average recharge to the aquifer from the catchment of 200–300 million m3 a−1. This led to a progressive decline in groundwater levels of 20–30 m and reduced flows in the Guadiana River. In turn, this resulted in a change in the hydrological functioning of the Tablas de Damiel to a recharge wetland rather than a discharge zone. The decline in the water table has resulted in the conversion of a net groundwater input into the wetland of 45 million m3 a−1, to a net outflow from the wetlands to groundwater of 33 million m3 a−1 (Llamas 1998). The ecological impact on the wetlands has been devastating. In places, the peats have become so dry that they have spontaneously combusted (Figure 28.6; Bromley 1996). The

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groundwater-licensed user. In practice, the situation has become much more complicated. Since the aquifer was legally declared ‘over-exploited’ in December 1994, and the Authority’s attempt to impose abstraction restrictions, farmers have probably drilled 8000–9000 new illegal water wells. Before this declaration, about 16 000 wells had been drilled (Acreman 2000a). A meeting of stakeholders took place in a local town (Villarrobledo) in May 1998. Graphs of groundwater level that were presented by farmers showed levels rising or constant, whilst those presented by the River Basin Authority showed groundwater level falling. There remains an urgent need for independent analysis of the data from all parties, dissemination of information on groundwater processes, current status and prediction for the future, together with improved stakeholder participation so that water users feel ownership of a catchment management plan. Only in this way will there be wider acceptance of legislation, abstraction licensing and other demand management measures. Fig. 28.6 Desiccation of the Tablas de Daimiel wetlands in central Spain, photographed in the 1960s (top) and the 1990s (bottom). (After Bromley 1996.)

CAT CHMEN T WAT ER QUALIT Y

Mancha Occidental aquifer was officially declared ‘over-exploited’ on 15 December 1994. An experimental plan to restore the wetland was approved in 1988 by the Spanish Government. This consists of three actions: (i) drilling of emergency pumping wells in the wetland; (ii) the transfer of up to 60 million m3 of water from another catchment to the Gigüela; and (iii) construction of a reservoir to supply the wetlands. This plan has concentrated on the symptoms and not the cause. There is still a need to encourage farmers, through better education, to use less water either by more efficient irrigation, by use of crops that require less water, or by a charging scheme that will discourage heavy use. The Spanish 1985 Water Act could theoretically solve the situation following the regulations to be applied in aquifers declared as ‘over-exploited’. For these aquifers, the Water Authority should prepare a Water Plan indicating the maximum amount of water available to each

All natural waters contain a variety of contaminants arising from erosion, leaching and weathering processes. To this natural contamination is added the pollutants arising from human sources. Wetlands are capable of assimilating a certain amount of pollution without serious effects because of the dilution and biological self-purification mechanisms that are present (see McCartney and Acreman, Chapter 17). However, if additional pollution occurs, the wetland system may be overloaded and degradation occurs. Heavy loads of nutrients such as phosphates and nitrates can cause excessive algal growth and depletion of oxygen with serious impacts on all aerobic organisms. Heavy metals and other pollutants can have direct impacts by killing plants, invertebrates, fish and birds. High copper levels, for example, can reduce the swimming ability of fish (Beaumont et al. 2000). The type of pollution that occurs is linked closely to water use and levels of socio-economic

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Hydrological Impacts in and around Wetlands development. Pollution remains a major threat to wetlands throughout the world. Industrialised countries have experienced a series of freshwater pollution problems involving domestic, industrial and agricultural wastes. Legislation and technology have been used to control particular sources of pollution as it has occurred. Today, pollution from industrial wastes is beginning to be curbed, but issues arising from non-point source pollution (i.e. acidification, organic micropollutants, nitrates etc.) are on the increase. Furthermore, illegal dumping of pollutants and industrial accidents still occur, sometimes with massive negative impacts for wetlands. In a study of 48 UK wetlands, 15% were threatened by water quality changes, largely because of agricultural activities on adjacent land (Gilman 1994). Furthermore, hydrochemical fluxes are often a key constraint to the restoration of degraded wetlands (Klotzli and Grootjans 2001). A study of the hydrochemistry of a reed bed constructed in southern England to provide bird habitat, found large diurnal variations in carbon dioxide pressure and dissolved oxygen, arising through photosynthesis and respiration. As the reed bed matured, organic litter built up because of both natural processes and management practices. This resulted in an increase in respiration and consequently an upward trend in average carbon dioxide pressure. It was anticipated that if the decomposition of organic matter continued to increase in the future, the associated decline in dissolved oxygen levels would make the reed bed an unsuitable habitat for some invertebrate species and fish (McCartney et al. 2003a). The Doñana National Park, Spain, is one of Europe’s most important wetlands, fed by the River Guadalquivir. Upstream is an important mining area producing zinc, lead, copper and silver. Crushed ore and chemicals, left behind after the metals have been removed, are deposited into a settling pond, held by an earth dam. In 1998 a 50 m-wide breach in the dam sent four million m3 of water and silt towards Doñana. Bulldozing of the river diverted much of the flood wave onto 2000 hectares of floodplain where silt with cadmium, mercury, arsenic and other heavy metals was deposited. Although the core area of

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the Park’s wetland was not polluted directly by the flood wave, metals leaked slowly into the groundwater. In addition, many of the birds resident in the park, and millions of migratory birds travelling to Africa (e.g. avocets and storks) fed on the adjacent floodplains and were killed by the pollutants (MacKenzie 1998). The River Tisza, tributary of the Danube in central Europe, has been dredged, straightened, shortened and impounded by flood protection embankments during engineering works that began in the 1880s (Csanyi et al. 2003). Work also disconnected meanders to form more than 150 oxbow lakes, many of which have become havens of biodiversity. As part of The River Tisza Project, funded by the European Commission, the water quality and ecological implications of increasing the hydraulic interaction between the main river channel and floodplain oxbows were investigated (Acreman et al. 2006). The research concluded that the ox-bow lakes would suffer significant degradation due to high nutrient levels and water quality in the main river would not be improved by their connection. In developing countries, the rapid growth of urban populations, particularly in South America and Asia, has outpaced the ability of governments to expand sewage and water infrastructure, and domestic waste is a major problem. Furthermore, these countries tend to have weaker laws, fewer resources with which to enforce them or merely an over-riding political will to ensure development even at the expense of the environment. The main water supply source for Harare (Zimbabwe’s capital), for example, is Lake Chivero. The City Council’s practice of dumping untreated wastewater into the lake led to high levels of organic matter that favoured the formation of carcinogenic micropollutants like chlorinated hydrocarbons. These are difficult to remove and local laboratories have only a limited capacity to monitor them. In 1997, the University of Harare found levels of mercury in the tap water to be 10 times higher than World Health Organisation guideline values. A government survey of Harare’s water supply in 1999 detected faecal contamination at some sampling points. On a technicality, the Council escaped prosecution by the Ministry of

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Water Development for continuing to pollute the lake (McCartney et al. 2000a). In response, the City Council had to spend nearly US$5 million to treat Harare’s drinking water (twice the amount spent in 1998).

S IT E M AN AGE M E N T Much of the concern about wetland degradation has been focused on the impacts of reduced flows caused by water abstraction, storage or diversion upstream. However, problems may also arise due to inappropriate, or lack of, site management. Today, the status of many wetlands is, to a greater or lesser extent, a consequence of human management. In many cases, without management these wetlands would have a very different ecological character. The transient nature of wetlands means that site management is necessary to conserve a range of wetland types in the landscape offering a variety of functions and values. In recent years, much emphasis has been placed on conserving smaller and more isolated wetlands within highly modified catchments, as their small size makes them less resilient and less able to withstand natural or human-induced changes. This may be vital to retain wetlands of particular historical interest, such as Wicken Fen (McCartney and de la Hera 2004). However, there is a pressing need to maintain and restore large wetlands, which can have significant beneficial functional impacts, such as balancing atmospheric carbon or reducing floods. Degraded wetlands are often revitalised by treating the symptoms rather than the cause, such as by pumping water on to a surface pond to provide flooding for wading birds, when natural inundation no longer takes place. Many managed systems have high conservation value, such as water meadows, which are a long established and accepted man-made system requiring continual management. Many problems result from cessation of traditional practices, such as reed cutting or dredging of the channel bed, which maintained highly valued aquatic systems.

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Extensive areas of former wetlands have artificial drainage systems, with surface ditches and plastic or clay soil pipes, typically 0.75–1 m deep. Britain (particularly England) is now one of the most extensively drained countries in the world. Between 1971 and 1980 about 10% of the farmland of England and Wales was drained (Robinson and Armstrong 1988). These drains lower the water table and soil water content in the fields as the water is removed to nearby watercourses. This reduces the potential for recharge to the regional groundwater table, although the supply of gravitydrained water from the upper metre of the soil will help to sustain river flows when normal soil water outflow from un-drained land has ceased. Current rates of new and renewal drainage are low, and the overall area of drained land is slowly declining. Peat cutting and gravel removal are two practices that have direct impacts on wetlands, as they remove large areas of a site. Extraction has the indirect effects of changing water levels and hydrological pathways and the rate of drainage to or from surrounding land or watercourses. Extraction often creates lakes, which in time may become important wildlife or recreation centres. For example, the Norfolk and Suffolk Broads, which now represent one of Britain’s finest wetlands, are flooded sites of former peat extraction dating back to medieval times (Lambert and Jennings 1953; Baker et al., Chapter 6). In the late 1980s and early 1990s extraction of gravel from Cassington pit, located about 2 km west of Oxford, induced groundwater level drawdowns in a number of adjacent water meadows. These ancient meadows known as Oxey Mead, Yarnton Mead and Pixey Mead lie on the northern and eastern banks of the River Thames, and are all classed as Sites of Special Scientific Interest (SSSIs). Studies were conducted to determine the drawdown likely to be encountered on the meads and to develop measures to minimise the impact (Bradford 1987). A groundwater model based on data collected over the previous 10 years was used to predict the impact of gravel extraction on groundwater levels under the meads. A number of measures to minimise the drawdown effects were simulated by the model.

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Hydrological Impacts in and around Wetlands Results showed the most effective option was to seal the working face of the pit closest to the meads with a blanket of Oxford Clay scooped up from the floor of the excavation. As the actively worked face moved along the boundary closest to the meads, a screen of clay was progressively plastered against the completed face left in its wake. A soil fill was then placed over the sealed section in order to minimise the extent and duration of groundwater drawdowns in the water meadows. Monitoring of water levels on the Meads during excavation subsequently confirmed the modelling predictions. The project demonstrated how it is possible to protect environmentally sensitive wetland sites from longterm damage by extractive industries using a carefully planned and staged programme of field investigations, modelling and monitoring. Changes to wetland vegetation may also significantly impact the hydrology of wetlands. In many wetlands, evaporation is the major pathway for exit of water. By changing the vegetation type, the evaporation rate can alter significantly. Phragmites reeds evaporate considerably more water than grass and some researchers claim the rate can be higher than that of open water (Hollis 1992).

M I T IG AT ION S T R AT E G IE S As this chapter has discussed, wetlands are altered because of changes in hydrological regime brought about by land conversion, agriculture, pollution, major engineering schemes and the demands of growing urban populations. Although there remains much debate about the best approach to reduce the impact of human intervention on freshwater ecosystems, there is growing recognition, at the highest political levels, that wide-ranging and integrated strategies are required to resolve complex problems. Towards this end, ecosystem management is proposed by advocates as the modern and preferred way of managing natural systems, including wetlands. The basic idea is that resource and landscape management is tailored to landscape conditions, processes and potential,

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so that desired features and processes of ecosystems are maintained and human needs are met in an optimal and sustainable way (Pirot and Meynell 1998; McCartney 2002). Mitigation strategies are intended to significantly reduce or eliminate defined threats or sources of wetland degradation. Two broad types of mitigation strategy are recognised. Remedial actions are those that try to prevent degradation at the end of the line (e.g. reducing nutrient levels in agricultural runoff by defining ‘buffer strips’ around farmland; see Blackwell, Chapter 19). Preventative actions are those that attempt to remove the cause of the problem (e.g. reducing the use of agricultural fertilisers and pesticides) rather than treating the symptoms. Some hydrological impacts, such as those resulting from climate change, may now be inevitable and little can be done in the short to medium term to reverse the changes. For other impacts, mitigation strategies have been developed to reverse or at least reduce the problem. In many cases the mitigation measure will be related directly to the cause. For example, the desiccation of a floodplain wetland due to retention of water in an upstream dam may be alleviated by release of managed floods from the dam (Acreman 1996a). In the late 1960s the Pongolapoort dam was constructed on the Pongolo River in north-east South Africa near its borders with Swaziland and Mozambique. The reservoir was filled in 1970 with a view to irrigating 40 000 hectares of agricultural land for white settlers, with no provision for hydropower generation. No assessments were undertaken of impacts of the impoundment on the downstream floodplain where 70 000 Tembe-Thonga people were dependent on recession agriculture, fishing and other wetland resources, nor on the biodiversity of the Ndumu game reserve. In the event, no settlers came to use the irrigation scheme. The dam changed the whole flooding regime of the river which led to crop failure on a massive scale. In 1978 plans were developed to make controlled releases of water from the dam to inundate the floodplain and rehabilitate the indigenous

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agricultural system and the wildlife. However, initial releases were made at the wrong time of the year and crops were either washed away or rotted (Poultney 1992). In 1987, the Department of Water Affairs and Forestry (DWAF) and the tribal authorities established water committees, representing five user groups: fishermen, livestock keepers, women, health workers (both new primary health care workers and traditional herbalists and diviners). They were given the mandate to decide when floodwaters should be released. Initially, these committees were very successful at implementing people’s views, and have since led to management of the river basin to the benefit of the floodplain users (Bruwer et al. 1996). However, since the mid-1990s, the effectiveness of the committees has declined, and in recent years, the participatory process has largely broken down. The failure of the water committees has been attributed to a number of factors, but is principally due to the lack of planning of natural resource use and development on the floodplain (Breen et al. 1998). Agriculture on the floodplain has grown at the expense of fisheries, tourism potential and biodiversity conservation, amongst others. More recently, the planning process has become increasingly complex as new stakeholders, including those upstream of the dam, have wanted to become involved in the decision-making process. Presently, DWAF continues to attempt to involve all interested and affected groups, but this is increasingly difficult (McCartney et al. 2003b). This highlights the difficulty of developing and implementing cooperative multi-stakeholder management systems. Some impacts may be impossible to reverse. Where groundwater-fed wetlands have been degraded by abstraction, resulting in reduced groundwater levels, it may not be possible to recover past levels if demand for pumping cannot be reduced. In these cases, alternative strategies may be developed, such as direct supply of water to a wetland from surface water. Where rivers have been separated from their floodplains by embankments there has been a loss of hydrological functions, such as flood storage (and thus increased flood risk downstream), and

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habitat and spawning grounds for fish (Welcomme 1979). Mitigation may involve removal of the embankments or re-engineering to allow local flood protection where needed, and controlled flooding of washland areas (Acreman 2000b). For example, serious flooding of cities in Germany and the Netherlands along the River Rhine during 1994 was made worse by the presence of embankments upstream. These had protected agricultural land on the floodplain, but prevented access by the river to natural floodwater storage. In response to these floods, embankments were removed and two large flood storage wetlands were created on the German bank of the Rhine as part of a programme to reduce flood damage downstream and restore degraded floodplains (Schrapp and Jans 2000). This type of environmental engineering marks a significant change in approach, which moves away from fighting against nature, towards working with it. Similar mitigation measures were proposed in the USA following the Mississippi floods of 1993. When embankments were over-topped, they prevented floodwater returning to the river, prolonging and intensifying the flooding. The US government agreed to review the entire basis of their flood defence strategy. This has prompted consideration of alternative solutions, such as exploiting the natural processes in the catchment, including wetland storage, which can play a role in regulating flood flows. These ideas have been given increased urgency by the New Orleans floods of 2005, which were possibly exacerbated by the loss of natural wetlands, and may occur more frequently in the future if climate change leads to more hydrological extremes as predicted. Improved water management within wetlands is also essential to make the best use of available water. Maintaining high water levels in wetland drainage channels is often used as a management tool, but this does not always achieve the ecological objectives. For example, improved drainage to increase agricultural production has been one of the major causes of degradation of the Somerset Levels and Moors. During the past 100 years, large investments, supported by government grants, have been made in pumping stations,

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Hydrological Impacts in and around Wetlands channel deepening and field drainage to reduce water levels, especially in the winter. The dryingout of the wetlands has led to a decline in many species, particularly wetland birds, and loss of the peat soils through oxidation. Internal Drainage Boards, dominated by farming interests have implemented water level management based on low winter ditch water levels, to allow drainage of floodwater, and high summer levels for stock watering and wet fencing. Essentially, this is the inverse of the natural situation, when much of the Levels and Moors were flooded during winter months, but became drier during the summer. The main tool for restoration of the Levels and Moors has been the raising of ditch water levels, in many areas maintained at around 200 mm below field level, to re-wet the wetlands. However, the low hydraulic conductivity of the soils has meant that only the field margins (up to 15 m from the ditches) benefit from raised ditch water levels (Armstrong et al. 1996), whilst the field centres remain dry. Achieving ecological objectives, such as restoring the numbers of breeding waders to the 1970 level, will require more widespread wetting, through inundation or reinstatement of the former shallow surface channels of the wetland. A potential undesired consequence of raised water levels in floodplain wetlands is that this may reduce floodwater storage volume; whilst floodplains may be important flood storage areas, storage is only available when the floodplain in dry. Studies on the North Drain of the Somerset Levels and Moors showed that current raised water levels reduced storage by 2% of the volume of the mean annual flood; however if all wetland ditches in the catchment were raised, this amounted to 84% of the mean annual flood volume (Acreman et al. 2006). In areas where peat has been extracted, leaving hollows that naturally fill with water, reed beds have been established. Some sites have been taken over by English Nature, the regulatory agency for nature conservation, or the Royal Society for the Protection of Birds, and reed establishment has been encouraged by planting and water level control to attract rare bird species such as bittern, marsh harrier and bearded tit. In the initial

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phases, sufficient water needs to be maintained to flood out terrestrial plant species such as willow, grasses and weeds, whilst not making it too deep at the reed planting stage early in the growing season; water levels are then raised progressively as the reed shoots grow (Gilman et al. 1998). Evaporation from reeds is 10% higher than from the original wet grassland (Acreman et al. 2003b) which has raised concerns from the Environment Agency, the government agency with responsibility for water resources, that a large increase in reed where water resources are stricken may lead to water shortages for other purposes, such as cattle watering, in the late summer. More research is required to confirm these figures. Technical solutions alone cannot halt the major loss of wetlands. Political will supported by targeted policies and enforced legislation are essential elements. The World Meteorological Organization has established an expert group to consider how ecological processes in wetlands and river systems can be integrated with flood management (WMO 2006). IUCN – The World Conservation Union – has established networks of experts to integrated catchment and wetland management, such as in Sahelian west Africa (Acreman 1996b). The World Bank has established an advisory panel on environmental flows to help ensure that sufficient water for rivers and wetlands is included within its water development projects, such as in Tanzania (Acreman et al. 2005). The importance of wetlands is increasingly being recognised globally and by December 2008 158 states had signed the International Convention on Wetlands initiated (in the town of Ramsar, Iran in 1971; Davis 1994). To be effective, however, national policies need to be implemented through local action. Member states of the European Union are implementing legislation to restrict surface and groundwater abstractions to meet the requirements of the Water Framework Directives in terms of maintaining ‘good status’ in all surface waters. Good status is defined in terms of the chemical, hydromorphological and ecological status of the water body. Although good status is not required for all wetlands under the Directive per se, wetlands linked to groundwater

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or other water bodies (such as lakes and rivers) and those designated (e.g. under the Habitats Directive) must achieve good ecological status. In the UK, Catchment Abstraction Management Strategies (CAMS) are being formulated to allocate sufficient water to maintain healthy rivers and wetlands (Environment Agency 2000) and to address the Water Framework Directive by defining maximum abstraction rates (Acreman et al. 2008) and ensuring adequate water is released from dams (Acreman et al. 2009b).

R E S E AR CH N E E DS The management of water resources is a key human challenge in this century. Science has a role to play in overcoming the numerous problems that confront us now and will do so in the future. Only during the last three decades has the value of wetlands, and the particular benefits that they provide, begun to be understood. It is now recognised that the formation, persistence, size and functions of wetlands are largely controlled by hydrological processes. Consequently, hydrology remains a key research discipline in attempts to ascertain the function of many wetlands, the extent to which they bring benefits to human society and the degree to which they may be used in a sustainable manner. Science offers the opportunity for better understanding of the inter-relationship between hydrological and biogeochemical cycles and provides the framework to link water resources planning with landscape and ecological planning. Policy development and decision-making need to be based on sound science. For example, legislation in Zimbabwe, dating from the 1920s and 1950s, restricts agricultural activity on headwater wetlands, known as dambos, because of the widely held perception that they regulate river flows. Recent research is questioning this assumption (Bullock and McCartney 1996; McCartney and Neal 1999) and rising population, in conjunction with efforts to increase food security and degradation of upland areas, means there is increasing pressure to change the law and utilise the wetlands for sustainable small-

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scale farming. Currently, the basis for making decisions about the extent and manner in which wetlands can be safely utilised for agriculture is weak. Greater understanding of the relationships between wetland utilisation, biophysical characteristics and socio-economic factors (driving wetland change) is essential for a full understanding of the causes and dynamics of wetland modification (McCartney and van Koppen 2004). Wetlands are worthy of scientific interest in relation to both process understanding and environmental management. As the global human population rises, there is increasing pressure on natural resources and wetland ecosystems. It is essential to develop a greater understanding of wetland dynamics in order to predict the impact of changes resulting from human activities. Fundamental research is required to improve understanding of basic processes, which are often complex and frequently counter-intuitive. Applied research is needed to produce tools for assessment of status and impacts and to develop sustainable management and restoration options. In particular research is required in the following areas: • Wetland processes – research to address factors that affect the type, location, size and functions of wetlands. • Wetland functions – research to determine the role wetlands play and the benefits they provide. • Human-induced stresses – research to improve ways of detecting or quantifying the effects of stress on wetlands and determining stress thresholds of wetlands. • Wetland restoration – research to develop tools and technologies to maintain, restore and construct wetlands. Knowledge gained from research can be used to provide an analytical perspective of problems, guide policy-making, and inform assessments of management interventions. However, hydrometric networks, fundamental to providing sound hydrological data, are globally in a state of decline and there is a particular paucity of records on the hydrological aspects of wetlands. It is often not possible, because of significant climatic and environmental differences, simply to transfer knowledge gained from research in one part of the world to another region. Furthermore,

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Hydrological Impacts in and around Wetlands although statistical methods and computer models have been developed as techniques for making assessments from inadequate data, there is still a need for real measurement of hydrological variable against which to calibrate and test models. There is thus a pressing need for improved data collection and the establishment of benchmark wetland monitoring sites.

I NTE G R AT IN G CAT CHM E N T AN D W E TLAN D S IT E M AN AGE M E NT The magnitude, frequency, duration and quality of water supplied to a wetland all have important controls over its character. Consequently, even subtle hydrological changes can have important impacts for ecological character, biodiversity and the processes that result in the products, functions and attributes that make wetlands so valuable. Hydrological change may result from alterations to climate, changes in catchment land-use, site management or through river engineering or groundwater abstraction. The single most important factor affecting wetlands in the future is likely to be demographic pressure, and related agricultural and urban land-use modification. Uplands surrounding wetlands have important controls on hydrology, so any management or restoration strategies require an integrated catchment approach. It is likely that, in the future, water for wetland conservation will become an increasingly important socio-political issue. Fundamental to successful management of wetlands and mitigation of negative human impacts is sound scientific understanding of wetland interaction within the landscape. It is essential that this is coupled with appropriate technology and combined with adequate policies, laws and incentives. Wetlands rely on water from the surrounding catchment, so they can only be managed successfully by taking a catchment perspective. This requires a catchment management plan that includes allocation of sufficient water to wetlands. Integrated catchment management is a manifestation of ecosystem management (Maltby et al. 1999) that recognises the need to

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manipulate and control our environment at the appropriate scale. Where surface water dominates, the river basin or catchment is the appropriate scale. However, frequently, an underlying aquifer does not coincide with the surface river basin. Thus, where groundwater plays a significant role, a group of basins overlying the aquifer may constitute the appropriate unit of water resource management (Acreman 1998). Within any plan, clear objectives for the wetland must be defined, including all products, functions and attributes that should be conserved. Thus, the catchment plan must be integrated with wetland site management plans; guidance on this provided by the Convention on Wetlands (Mackay et al. 2009). Although wetland objectives are ultimately a matter of societal choice, science can help define what is possible, how it can be achieved and whether it is sustainable. A key prerequisite is the establishment of the water needs of various species in the wetlands, which may include birds, plants and invertebrates. In many cases too much water at the wrong time can be as damaging to the wetland as too little water. The definition of water needs must encompass magnitude, frequency, duration and quality of water. In addition, an implementation plan is required to ensure that the objectives of the catchment management plan are met. In many catchments, there is inadequate water to easily satisfy all demands. Allocation of sufficient water to wetlands may require the operation of dams, pumps, sluice gates, penning boards and other devices. Therefore, the implementation plan needs to include rules for the operation of infrastructure. A key element of the plan is adequate monitoring to measure compliance. In some cases hydrological forecasting may be required, based on computer models, to predict the consequences for water allocations to the wetland of water management in the catchment, such as water stored in dams. These models may also be used to examine the consequences of future water management scenarios or possible climate change. Contemporary scientific tools can assist greatly in the management of wetlands. For example, a study of the Usangu wetlands in Tanzania was designed to assess the impact of irrigation on

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the hydrological regime of the wetlands, and to evaluate the magnitude of inflows required to maintain flows downstream of the wetlands. The study used hydrological data, in conjunction with remote sensing and GIS techniques, to provide insights into the dynamics of change and the spatial response of the wetlands. A water balance model was developed to simulate the major components of the water budget and derive estimates of inflows required to ensure a range of possible downstream flows (Kashaigili et al. 2006). Effective wetland management relies partly on existence of data and appropriate technical tools to recognise, assess and mitigate hydrological impacts. Of equal importance is the need for a range of supporting legal and institutional mechanisms, and awareness of wetland issues including: • institutions with appropriate technical expertise to manage wetlands effectively • national policies to promote wetland conservation and their wise use • appropriate legislation to protect wetlands • coordinating mechanisms to allow stakeholders (including government departments, local authorities, NGOs and local residents) to meet and exchange information • awareness amongst planners, decision-makers and the general public of the functions and values of wetlands. The International Convention on Wetlands produced a set of guidelines for the allocation and management of water for maintaining the ecological functions of wetlands (Acreman et al. 2002). These guidelines recognise the fundamental need for sound science and its utilisation to support decision-making. The principles in the guidelines are: • sustainability as a goal; • equity in participation and decision-making factors; • clarity of process; • credibility of science; • flexibility of management; • accountability for decisions; • transparency in implementation. It is evident that sound scientific understanding of physical, chemical and biological processes is

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essential at a range of scales from micro to global for the successful management of wetlands. A key scale is the catchment, as this is the fundamental unit for management of water resources that determines the allocation of water between wetlands and other uses and, ultimately, the viability of a site as a wetland. R EFER EN CES Abramovitz J.N. 1997. Valuing nature’s services. In: State of the World. Worldwatch Institute Report on Progress towards a Sustainable Society, The World Watch Institute, Washington, DC, pp. 92–114. Acreman M.C. 1996a. Environmental effects of hydroelectric power generation in Africa and the potential for artificial floods. Water and Environmental Management 10(6), 429–434. Acreman M.C. 1996b. The IUCN Sahelian floodplain initiative – networking to build capacity to manage Sahelian floodplain resources sustainably. International Journal of Water Resources Development 12(4), 429–436. Acreman M.C. 1998. Principles of water management for people and the environment. In: de Shirbinin A. and Dompka V. (editors), Water and Population Dynamics. American Association for the Advancement of Science, Washington, DC, 321 pp. Acreman M.C. 2000a. Guidelines for the Sustainable Management of Groundwater-Fed Catchments. Report to the European Union ENV4-CT 95-0186 – Groundwater and River Resources Programme on a European Scale (GRAPES), Institute of Hydrology, Wallingford. Acreman M.C. 2000b. Wetlands and Hydrology. Mediterranean Wetlands Programme Publication Number 10, Tour du Valat, France, p. 109. Acreman M.C. 2002. Case Studies of Managed Flood Releases. Environmental Flow Assessment Part III. World Bank Water Resources and Environmental Management Best Practice Brief No. 8, World Bank, Washington, DC. Acreman M.C. 2005. Guidelines for the Management of Groundwater to Maintain Wetland Ecological Character. Report to the International Convention on Wetlands (Ramsar). Centre for Ecology and Hydrology, Wallingford, 25 pp. Acreman M.C. 2006. Review of the Restoration of Hydrological Functioning of Floodplain Wetlands in the Inner Niger Delta, Mali. Report to IUCN – The

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Hydrological Impacts in and around Wetlands World Conservation Union, Mali. Centre for Ecology and Hydrology, Wallingford, 27 pp. Acreman M.C., Aldrick J., Binnie C., Black A.R., Cowx I., Dawson F.H., Dunbar M.J., Extence C., Hannaford J., Harby A. et al. 2009b. Environmental flows from dams; the Water Framework Directive Engineering Sustainability 162, (ESI), 13–22. Acreman M.C., Barbier E.B., Birley M., Campbell K., Farquharson F.A.K., Hodgson N., Lazenby J., McCartney M.P., Morton J., Smith D. et al. 2000. Managed Flood Releases – Issues and Guidance. Report to DFID and the World Commission on Dams. Centre for Ecology and Hydrology, Wallingford. Acreman M.C., Blake J.R., Booker D.J., Harding R.J., Reynard N., Mountford J.O. and Stratford C.J. 2009a. A simple framework for evaluating regional wetland ecohydrological response to climate change with case studies from Great Britain. Ecohydrology 2, 1–17. Acreman M.C., Booker D.J., Riddington R. 2003a. Hydrological impacts of floodplain restoration: a case study of the river Cherwell, UK. Hydrology and Earth System Sciences 7(1), 75–86. Acreman M.C., Dunbar M.J., Hannaford J., Wood P.J., Holmes N.J., Cowx I., Noble R., Mountford J.O., King J., Black A. et al. 2008. Developing environmental standards for abstractions from UK rivers to implement the Water Framework Directive. Hydrological Sciences Journal 53(6), 1105–1120. Acreman M.C., Fisher J., Stratford C.J., Mould D.J. and Mountford J.O. 2006. Hydrological science and wetland restoration: case studies from Europe. Hydrology and Earth System Sciences 11(1), 158–169. Acreman M.C., Harding R.J., Lloyd C.R. and McNeil D.D. 2003b. Evaporation characteristics of wetlands; experience from a wet grassland and a reedbed using eddy correlation measurements. Hydrology and Earth System Science 7(1), 11–22. Acreman M.C. and José P. 2000. Wetlands. In: Acreman M.C. (editor), Hydrology of the UK – A Study of Change. Routledge, London, pp. 204–224. Acreman M.C., King J., Hirji R., Sarunday W. and Mutayoba W. 2005. Capacity building to undertake environmental flow assessments in Tanzania. In: Proceedings of the International Conference on River Basin Management, Morogorro, Tanzania, March 2005. Sokoine University, Morogorro. Acreman M.C. and Lahmann E. 1995. Hydrological management and protected areas. PARKS. The International Journal for Protected Area Managers 5(2), 1–5. Acreman M.C., Mackay H. and Cowan G. 2002. Guidelines for Allocation and Management of Water

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29

Biotic Pressures and Their Effects on Wetland Functioning C. MAX FI NLAYSO N

Institue for Land, Water and Society, Charles Sturt University, Albury, Australia

IN T R O D U CT ION The importance of wetland dysfunction as a consequence of biotic pressures exerted by invasive species has not always been as widely appreciated as may now be assumed. Until recently, the problem of invasive species in wetlands was seen primarily as a problem for developed countries; a situation that persisted despite the paradoxical occurrence of well documented cases of invasive species in African wetlands and lakes (e.g. Salvinia molesta and Nile perch – Lates nilotica). The reasons for this lack of appreciation are not clear, although they probably included: (i) insufficient public and institutional awareness of the problems; (ii) insufficient information about the species and ways of controlling them; or (iii) insufficient capacity to collect information or implement control measures. However, recent initiatives (e.g. the Global Invasive Species Programme, www.gisp.org) have demonstrated that wetland dysfunction caused by invasive species is now much more of a concern globally (Pittock et al. 1999; McNeeley et al. 2001). The importance of invasive species in wetlands is examined in this chapter. This includes consideration of the terms and concepts currently in use, global initiatives, features of invasive species, and ways of managing them. While specific cases of adverse ecological change caused

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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by invasive species are presented, analysis is still limited by the scarcity of quantitative studies. This situation has arisen despite extensive efforts to manage or control these species. The tendency to control species without initially ascertaining the extent of the problem is not new; it is ongoing and hinders the effective management of at least some invasive and alien species (Rea and Storrs 1999; Mooney 2001).

CON CEPT S AN D T ER MS The concepts and terms that have been used when discussing biotic dysfunction of wetlands have caused confusion at times. Recent initiatives have not only raised the profile of biotic disturbance in wetlands due to invasive species, but have also elaborated the concepts of alien species and population variability in relation to natural change and the environmental tolerances of many species. The latter may need further consideration. Not only can environmental conditions influence the invasive potential of species, but they can also, at least partly, determine the movement of species beyond what may be regarded as their ‘normal’ ranges, and hence when they can be considered as invasive or alien. Simply moving beyond a ‘normal’ range does not always imply that a species should be considered alien or invasive (in an adverse sense). This is well illustrated by species of water birds that can move far beyond their ‘normal’ range in response to adverse weather, such as the effect of cold climate or drought on

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water bird movements (Ridgill and Fox 1990; Kingsford 1995), and not be treated as alien or invasive species. In fact, in some instances these movements may reflect the full extent of their range or, if the range is extended, may be seen as favourable by some sectors of society. It is contended that societal attitudes play a major role in determining whether species that may be out of their ‘normal’ range are considered invasive or alien. Coupled with variability in population movements, we are faced also with increasing movements of species due to direct human interventions, such as deliberate transport from one region to another, or changes in habitat suitability (associated with habitat degradation, but also with so-called ‘beautification’ and the creation of non-natural habitats). Such changes are critically important in enabling displaced or translocated species to gain a ‘foot-hold’ in wetland habitats where they previously did not occur (Rea and Storrs 1999). Further, although not addressed specifically here, there is increasing concern that genetically modified organisms could become invasive, and compound existing problems with alien species (Beringer 2000; McNeeley et al. 2001). In an attempt to standardise the terms used to describe invasive species, whether alien or native, the following definitions were accepted at a Global Biodiversity Forum (GBF13) workshop on alien and invasive species (Pittock et al. 1999): • An alien species is a species, subspecies or lower taxon, occurring as a result of human agency in an area or ecosystem in which it is not native; • An invasive species is an alien species, which colonises natural or semi-natural ecosystems, is an agent of change, and threatens native biological diversity; • The terms pests or weeds are sometimes used as synonyms for invasives (i.e. alien species threatening native biodiversity), but at other times are used when referring to alien species threatening agriculture or forestry. The Global Invasive Species Programme adopted further definitions that extended some of

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the concepts agreed by GBF13 (McNeeley et al. 2001): • An alien species (synonyms: non-native, non-indigenous, foreign, exotic) is a species, subspecies or lower taxon, introduced outside its normal past or present distribution; includes any part: gametes, seeds, eggs or propagules of such species that might survive and subsequently reproduce; • An invasive alien species is an alien species whose establishment and spread threaten ecosystems, habitats or species, with economic or environmental harm; • A pest is any species, strain or biotype of plant, animal or pathogenic agent injurious to plants or plant products (as agreed by the International Plant Protection Convention); • A weed (synonyms: plant pest, harmful species; problem plant) is a plant (not necessarily alien) that grows in sites where it is not wanted and has detectable negative economic or environmental effects; alien weeds are invasive alien species. Given that many pest species have been considered traditionally within an economic context, for example, pests of agricultural enterprise, it is necessary to promote greater consideration of pest species in a context that incorporates broader societal attitudes towards environmental change. In this context the following definition has been adapted from the definition of an environmental weed provided by Humphries et al. (1991): Environmental pests are those species that invade native communities or ecosystems and are undesirable from an ecological perspective, but not necessarily an economic one.

The concepts expressed by this definition are reflective of a greater emphasis placed on environmental or ecological values within the context of sustainable development of wetlands. In the text that follows, the emphasis is placed on invasive species, whether alien or not, that adversely affect the ecological character of wetlands (i.e. results in biotic dysfunction) with or without direct economic affects.

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Effects of Biotic Pressures on Wetland Functioning G L O B AL IN IT IAT IV E S In response to the increasing recognition of the problems caused by invasive species in many ecosystems, the Global Invasive Species Programme (GISP) was established in 1996 with the following objectives: • to assemble and make available best practices for the prevention and management of invasive alien species; and • to stimulate the development of new tools in science, policy, information and education for addressing these problems. To achieve these objectives, an international network of experts was assembled, and an initial 3-year work plan established with specific practical outputs (Waage 2001). Key outcomes included a toolkit for managing invasive alien species (Wittenburg and Cock 2001) covering strategies and policies, methods for prevention, risk assessment, early detection and management approaches. A further major outcome was a global strategy for invasive alien species (McNeeley et al. 2001) that highlighted the dimensions of the problem and outlined a framework for mounting a global-scale response. The strategy drew attention to the toolkit and outlined activities to implement the following five initiatives: • Global access to information on threats and their prevention and management; • Directed action at key pathways of introduction, through public and private sector cooperation; • Acceleration of critical research and its dissemination; • Awareness-raising and support to policy development; • Building cooperation between institutions towards a global biosecurity platform to mitigate the threat of invasive alien species. GISP has continued to support wider awareness and information programmes on invasive species, with recent reports being produced on alien invasive species in Asia (Matthews 2004) and Africa (Matthews and Brand 2004). Many of the species illustrated in these reports have invaded wetlands.

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Whilst GISP has focused much of its attention on the invasive species initiatives of the Convention on Biological Diversity (www. biodiv.org), other initiatives have more directly focused attention on invasive species in wetlands. The concepts outlined in GBF13 were adopted in a resolution by the Ramsar Wetlands Convention on invasive species and wetlands (http://www. ramsar.org/key_res_vii.14e.pdf; Ramsar 1999). The resolution promoted a number of concepts and actions, including: • Consideration of the environmental, economic and social impact of invasive species; • Preparation of wetland-specific guidelines for identifying and establishing priorities for action and management of troublesome alien species; • Provision of guidance on legislation or other best practice management approaches that incorporate risk assessment; • Preparation of inventories of alien species in wetlands to assess them and identify management options. Wetland-specific guidelines were drafted by the Scientific and Technical Review Panel of the Ramsar Convention but, owing to wider political considerations, they have not been adopted by the Convention. Given the absence of wetland-specific guidelines, attention is drawn to the generically applicable guidelines produced by the Species Survival Commission of the World Conservation Union (http://intranet.iucn.org/webfiles/doc/SSC/ SSCwebsite/Policy_statements/IUCN_ Guidelines_for_the_Prevention_of_Biodiversity_ Loss_caused_by_Alien_Invasive_Species.pdf). These cover: definitions and terms; understanding and awareness; prevention and introductions; eradication and control; re-introductions; knowledge and research and laws and institutions (see Internet address at end).

POPULAT ION GR OWT H AN D MIGR AT ION In considering invasion of wetlands, we need to consider the balances and checks that control, or at least influence, the population growth and

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movement of specific taxa. Many species are migratory, and regularly move from one location to another in response to external stimuli such as climate or resource availability. Further, many populations fluctuate in response to similar stimuli. Thus, when deciding whether a species is invasive, we need to assess the nature of population fluctuations and movements within its ‘normal’ range. Extremes of climate can cause species to move outside of their normal range. However, our ability to assess such changes is limited by the quality of population and habitat or range data. The difficulty in obtaining such species inventory information has been shown recently for migratory water birds, which have been the subject of much survey and assessment (Delany et al. 1999; Wetlands International 2002). The assessment of populations and native ranges can be further complicated by species that undergo intermittent or even episodic fluctuations, generally in response to climate patterns. Whilst biologists or land managers may consider that they have sufficient broad understanding of such fluctuations, we are yet to witness, let alone assess and monitor, the full impact on species and habitats of global-scale climate change and sea level rise (RSPB 1997; Gitay et al. 2002; van Dam et al. 2002a). The extent of potential displacement or loss of wetland habitats due to climate change is still a contentious issue. Finlayson (1999) used a model scenario from northern Australia (Eliot et al. 1999) to argue that we need to consider potential changes due to climate change as part of all management and sustainable development considerations for wetlands. We cannot assign with certainty a ‘native range’ to all species, especially those that are highly mobile or respond to climate variability. However, whilst this could prove problematic for some species, there is no doubt that many others have been removed from their native range and established in non-native habitats. Examples are the pan-tropical weeds salvinia (Salvinia molesta) and water hyacinth (Eichhornia crassipes) that originated in South America, but which are now widely distributed across the tropics. Animals that have spread and disrupted the

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wetland habitats they have invaded include cane toad (Bufo marinus), European domestic pig (Sus scrofa) and carp (Cyprinus carpio). While there are some obvious examples of species that have established outside their native range, it should not be assumed that all newly established species are alien and have been ‘transplanted’ by human activities. As noted above, recent concerns over global climate change and variability provide a scenario where it may no longer be possible to attribute the occurrence of ‘new’ species to natural fluctuations versus human activity. Under such a scenario the importance of ecological information and knowledge about species and their distribution is of paramount importance. Understanding the interactions between population growth and invasion has been a challenge faced by managers of wetland pests for many years. Species such as the water hyacinth (Eichhornia crassipes Figure 29.1) and Canadian pondweed (Elodea canadensis) have spread around the globe (Sculthorpe 1967; Gopal 1987) with massive changes in population levels apparently uninfluenced by control or management programmes. This prompted the warning by Mitchell (1978) that managers need to consider whether effective action could be implemented or whether

Fig. 29.1 Water hyacinth (Eichhornia crassipes) is a widespread floating weed that can cover open water bodies, blocking light penetration and resulting in deoxygenation. Extensive infestations can disrupt boat transport. (Photo by author.)

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Effects of Biotic Pressures on Wetland Functioning they should be inactive, and refrain from taking non-effective actions. The latter illustrates an aspect of wetland management that seems to be ingrained in many practitioners: there is an impatient need to take immediate action even when success is unlikely (‘at least we tried!’). This is well-illustrated in Australia where there have been repeated interventions against Salvinia molesta, but only rare successes (Finlayson and Mitchell 1982; Finlayson et al. 1994; Storrs and Julien 1996). Further, there might be occasions where the incumbent generation of wetland or pest managers will criticise their predecessors for not taking action when there was (supposedly) a chance to achieve the desired results. Rea and Storrs (1999) discuss the penchant for wetland managers to tackle the problem rather than tackle the underlying reasons that caused the problem. The example of Canadian pondweed is illustrative of the dilemma faced by wetland managers (see Sculthorpe 1967; Mitchell 1978). It is the first documented example of the explosive growth of an aquatic weed; it originated in North America and invaded the waterways of Europe in the late nineteenth century. Immediately after it invaded a locality it grew rapidly, reproducing vegetatively, and reaching a maximum population density within a period of a few months to 4 years. It maintained this population for up to 5 years and then declined to levels that were not considered a nuisance. The exact reasons for the rapid increase and subsequent decline were not determined. It does not take a great deal of imagination to discern the level of panic that could prevail if ‘do nothing’ was proposed for other invasive species with explosive population increases. The wisdom of taking action is dependent on a number of factors, including the likelihood of success and the likely consequences if the species continues to spread rapidly.

M E T H O D S O F IN V AS ION Pittock et al. (1999) introduced the issue of alien and invasive species in wetlands with the following statement ‘Increasing travel, globalisation of

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the economy, improper management practices, agricultural, and other commercial programmes have accidentally or deliberately transported non-indigenous species to wetland ecosystems, where they have become harmful invasives (usually categorised as “pests” or “weeds”)’. Further, it was noted in the same forum (GBF13) that ‘… responsibility for invasions lies with human action in the form of deliberate or unintentional transport of species, coupled with poor environmental management’. Thus, it is considered generally that human endeavour either enables invasive species to be transported from one locality to another or, importantly, results in habitat conditions that enable introduced or native species to thrive and become a nuisance. Williamson (1996) introduced the principle that 10% of alien species introduced into a region will appear in the ‘wild’, 10% of these will establish and 10% of those that establish are invasive. Manchester and Bullock (2000) queried the figures used to establish this principle, but acknowledged that only a small proportion of introduced species are likely to flourish and become a serious problem. These species obviously have specific traits, which enable them to reproduce or establish in new habitats or areas within the native range, which have been altered or degraded. Management responses need to consider both the characteristics of the particular species and the habitats that are prone to invasion. Thus, the manner in which propagules are transported and established in specific habitats needs to be considered. Many wetlands are prone to invasion by propagules of species that are carried by water or by highly mobile or migratory animals, as possibly occurs for many wetland plants. Further, low species richness in some wetland habitats may enable propagules to establish readily (Figure 29.2). This was the situation in northern Australia, where the removal of large numbers of feral Asian water buffalo (Bubalus bubalis; Figure 29.3) allowed the alien weed Brachiaria (Urochloa) mutica to establish (Cowie 1996). Similar consequences of reduced grazing pressure are described by Gopal (Chapter 3). However, the inter-relationships between species diversity

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Fig. 29.2 Paragrass (Urochloa mutica) is a perennial stoloniferous grass from Africa that has been widely spread throughout the tropics and subtropics as a pasture species and for controlling riverbank erosion. In northern Australia it is now considered a pest species whereas in parts of India, for example, it is cultivated along riverbanks and floodplains for cattle fodder. (Photo by author.)

Fig. 29.3 The Asian water buffalo (Bubalus bubalis) has long been associated with people and wetlands in Asia and has been introduced to tropical and subtropical wetlands elsewhere. In northern Australia high densities and over-grazing resulted in large numbers being harvested and removed from floodplain wetlands. (Photo by author.)

and habitat characteristics of wetlands, and invasion by alien species, deserve further investigation. van Dam et al. (1999) proposed the use of formalised risk assessment procedures, such as those accepted by the Ramsar Wetlands Convention (http://www.ramsar.org/res/key_res_vii.10e.pdf). This involves six steps: • identification of the problem; • identification of the effects; • identification of the extent of the problem; • identification of the risk; • risk management and reduction; and • monitoring. Examples of the application of risk assessment to wetland invasive species are presented by van Dam et al. 2002b and Walden et al. 2004. The outcomes of these assessments have been used to provide advice for risk management and monitoring. Risk assessments such as those mentioned above have principally addressed the ecological reasons for change in wetlands: Finlayson and Rea (1999) contended that the non-ecological reasons for wetland loss should be considered alongside

the ecological reasons. Their generalised concepts of ‘ecological’ and ‘non-ecological’ reasons are listed below: • Ecological reasons for invasion of wetlands. Many species have features that enable them to rapidly invade and establish in wetlands, especially where the habitats have been disturbed, or even when natural cycles enable them to become rapidly distributed and to reproduce. • Non-ecological reasons for invasion of wetlands. Land use and regulation are key factors behind the invasion of wetlands by pest species. This relates to direct disturbance of the wetland and to trade and transport that allows species to be moved rapidly from one location to another. Whilst many species have features that enable them to take advantage of changed ecological conditions (Newsome and Noble 1986; Barrett and Richardson 1986), there are as many nonecological reasons that result in the movement of species from their native range or in an imbalance in ecological conditions that would otherwise limit population levels.

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Effects of Biotic Pressures on Wetland Functioning M E T HODS OF CON T R OL There are many specific control measures recommended for individual invasive species, some of which are outlined in the case studies given below. As there are many individual and well-documented approaches to controlling invasive species, it is sufficient to say that, in many instances, single solutions are not as effective as integrated and multi-faceted solutions. For many years, generic principles for addressing invasive and pest species have been available, and these are still valid. Unfortunately, acceptance of or adherence to such principles seem to be weak or lacking (Finlayson and Rea 1999). As an example, Mitchell (1978) proposed that aquatic weeds within Australia were addressed in a series of steps that included: • assessment and monitoring of infestations using systematic and standardised methods; • elimination of potentially dangerous infestations at the earliest opportunity; • prevention through quarantine and import regulations; • consideration of potential problems during land planning exercises; • increased awareness and public education; • greater cooperation between government sectors and private developers; • increased knowledge of the causes of weed problems and possible control methods. At a conceptual level, prevention, quarantine and early intervention are attractive components of any management regime. However, these are fraught with difficulties in terms of identification, surveillance and actual control of any infestations, and need underpinning by awareness, education and knowledge across many sectors of society and government. Difficulties with implementing such multi-faceted approaches have occurred as a consequence of conflicts with the proponents of economic development, both locally and globally, and also as a consequence of increased volumes of trade and transport of many goods, again, locally and globally. Given these factors, the management of invasive and pest species should not be considered in isolation. Every

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effort should be made to ensure that collaborative approaches for controlling invasive species are included in all strategies for sustainable development (within appropriate geo-political realms). Economic realities also include local issues, such as the cost of herbicides and labour, and are also influenced by cultural traditions and sensitivities (e.g. emotive responses to the shooting of mammals from helicopters) that usually need to be considered on a case-by-case basis. In general, weed species are often tackled through a combination of physical, chemical or biological control techniques. Early intervention and the control of satellite infestations are recommended methods, with consistent and ongoing surveillance and rehabilitation required after control (Cook et al. 1996). In recent years, biological control methods have been promoted as an ecologically sound approach to managing pest species, and some successes have received publicity. However, over-optimism that biological control will readily provide a solution should be tempered with the realisation that such programmes can take many years to develop and may not be effective. Further, where such agents are employed, it may not always be easy to relate ‘cause and effect’ within complex ecosystems. The control of animal pests is based on similar premises as weeds, with harvesting of large species being effective and a popular method with managers, even though shooting can meet with social opposition. Again, however, prevention and early intervention, backed by containment where possible, are recommended methods. Smaller vertebrate and invertebrate species provide another level of problem, as early detection may not occur and surveillance is difficult. The introduction of disease bearing mosquitoes is one example (Clout and Lowe 2000) while the spread of many species of fish is another (Kolar and Lodge 2000). In all control or management programmes, it is necessary to support field assessments and actions by well targeted monitoring and research. The former should be based on clear objectives, be constantly reviewed, and cover both the target

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species and the control actions. Research needs to be targeted and integrated with the management programme and should not be seen as a separate or luxury exercise. Further, the management programme should not stop when control is affected; there may be a resurgence of the problem, and possibly a need for rehabilitation of the locality or habitat in question. Lack of funding for postcontrol management can undo all the benefits of a control programme, and illustrates the futility of not ensuring that funding and management

are supported by a long-term commitment by all responsible sectors.

CASE ST UDIES In order to illustrate the problems caused by invasive species and the means of managing them, case studies are presented by researchers closely involved with investigations into these species and the problems they cause.

Case study 1 Nile Perch invasion in Lake Victoria Geoffrey Howard IUCN, Nairobi, Kenya Nile perch (Lates niloticus) is a large predatory fish of the family Centropomidae that can grow to 1.8 m in length and weigh as much as 200 kg (Figure 29.4). Lates niloticus is native to the White Nile River system (in Uganda, Sudan, Ethiopia and Egypt) including Lake Albert that it shares with another species, L. macropthalmus.

Fig. 29.4 Specimens of Nile perch (Lates niloticus) less than 1 m long. The species had been introduced to Lake Victoria to boost fisheries. Efforts to introduce it further afield to Australia have been resisted in favour of developing fisheries using the closely related native barramundi (Lates calcarifer). (Photo by M. McCartney.)

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L. niloticus (together with L. longispinis) is present naturally in Lake Turkana (Kenya) that had ancient connections to the Nile system. Nile perch was not thought to be present in Lake Victoria and Lake Kyoga until introduced from Lakes Albert and Turkana in the 1950s and 1960s (Lowe-McConnel 1987). Lake Victoria (and its satellite lakes such as Nabugabo, Bisina, Olpeta and Kanyaboli) was famous for its diverse fish fauna of around 350 species – 300 of them being endemic cichlids of the subfamily Haplochromiinae. These small fish occupied many niches in the lake and its wetlands and exploited many and varied sources of food (Witte et al. 1995). While the haplochromines were mostly too small for human food, there was a multi-species fishery of the other, larger species, which supplied many riparian communities with protein and supported commercial operations. The Nile perch was introduced to Lake Victoria to boost this fishery, but was hardly seen in catches until the late 1970s (Goldschmidt et al. 1993). Nile perch catch statistics from Lake Victoria then tell of a dramatic increase in this predator. In 1978 around 1000 tonnes of Nile perch were caught in Kenyan waters of the lake; by 1981 this had risen to 23 000 tonnes and by 1985

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Effects of Biotic Pressures on Wetland Functioning to 50 000 tonnes. The Kenyan catch reached a peak of 123 000 tonnes in 1991. Lake-wide the same story is told, with the entire lake yielding less than 25 000 tonnes in 1981 but producing 363 000 tonnes of Nile perch by 1993. At the same time, the total catch of all species rose from around 100 000 tonnes in 1979 to about 500 000 tonnes in 1989. The proportion of Nile perch in the fish catch rose from less than 0.1% in 1974 (Goldschmidt et al. 1993) to more than 50% 20 years later. This is the basis for calling Nile perch in Lake Victoria an (alien) invasive species. As the population of Nile perch increased, the diversity and mass of many other fish populations in the lake decreased, at an equally dramatic rate. Haplochromines made up 90% of the weight of experimental trawl catches in the northern region of the Lake Victoria in the 1970s and early 1980s. By 1985, however, this had fallen to 7%, while Nile perch had increased from a negligible presence in the 1970s to 90% by 1985 (Ogutu-Ohwayo 1999). Many authors have also reported that during the same period, the main lake lost many of its species of haplochromines; even as many as 50% of the endemic species. After the depletion of the small cichlids, the lake became dominated by three species, Nile perch, the introduced Tilapia nilotica and a native cyprinid (sardine) Rastrineobola argentea. To date these are the three main species in the fishery, to the exclusion of most of the native palatable and nonpalatable species. Other changes occurred in the ecology of Lake Victoria at the same time, and are attributable to overuse of lake resources and advancing eutrophication from the watershed. The advent of the invasive alien water hyacinth (Eichhornia crassipes) has brought about further degradation of the lake so that it is difficult to attribute alterations in biodiversity solely to the Nile perch invasion. Nevertheless, there is little doubt that the invasion of Nile perch has resulted in the reduction of many other fish populations and the extinction of some endemic species, and

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is likely to have affected other organisms and habitats within the lake. At the same time, the immense increase in Nile perch and the expansion of the fishery tonnage brought great benefit to a small proportion of the riparian peoples on Lake Victoria. A vast commercial fishery has developed, which harvests the Nile perch, processes the fish in (mostly) lakeside factories and exports fresh and frozen fish fillets and ‘steaks’ to Europe, North America and other foreign locations. This has brought great benefit to some and has increased the foreign exchange earnings of the lake to the level of many millions of US dollars per year. The same process, however, has deprived most riparian communities of access to essential fish protein, as Nile perch brings a higher price than local people can afford. It has also removed many subsistence fishers from their traditional roles in the fisheries of the lake. A direct result of this is that many traditional fishers of Lake Victoria have been forced to use unsustainable fishing methods to harvest enough fish for their survival, further depleting the biodiversity resources of the lake. There are a number of probable causes of the dramatic population increase in the invasive Nile perch: • Lack of other wide-ranging predators (competitors); • A fast growth rate and reproductive potential; • A great range of body size during development which permits exploitation of various habitats in the lake; • Changes in the lake ecosystem resulting from human activities (e.g. eutrophication, increase in the depth and area of the anoxic layer, changes in phytoplankton and herbivorous zooplankton (Ogutu-Ohwayo 1999)) which negatively affect other fish species but have little impact on Nile perch; • Adaptability of the perch to different sources of food (haplochromines, then other native cichlids, introduced cichlids, the sardine

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R. argentea, the native prawn Caridina nilotica and its own young); • Possible hybridisation among the introduced species of Lates and consequent ‘hybrid vigour’ from elements of the three species from lakes Albert and Turkana (Witte et al. 1995). Can the Nile perch expansion in Lake Victoria be controlled? This is unlikely and undesirable for those making vast profits from the export industry. It is still not clear whether the Nile

perch population and the Lake Victoria fish community structure has reached (or is anywhere near) stability. How can 50% (or more) of the biomass of fish in a lake be made up of one species of predator? Perhaps the most we can do is to monitor this phenomenon, document it, learn from it and make sure that no other lakes (particularly the great lakes) of Africa are assailed by introductions of alien species with only fisheries’ profits in mind.

Case study 2 Common Carp Cyprinus carpio L. Jane Roberts Consulting Biologist, Canberra, Australia Common carp, Cyprinus carpio L., is one of the most widely distributed fish species in the world and has considerable economic and ecological importance. Carp have been grown for food for centuries in parts of Asia and Europe and are still an important food source, despite political and economic change. They are bred in Japan for certain colour variants and these ornamental ‘koi’ carp provide a valuable international trade in live fish. In the last 200 years, feral populations of domesticated forms have established around the world in a range of climates from cool temperate to semi-arid and subtropical. These feral carp are found in natural, modified and constructed habitats; in a range of water qualities and water regimes; and in lotic and lentic conditions. Examples of feral carp habitat include riverine and non-riverine wetlands, regulated rivers, reservoirs, weir-pools, urban lakes, irrigation channels and systems, and constructed wetlands. In Australia, New Zealand, United States, South Africa and Canada, the feeding behaviour and abundance of feral carp makes theman environmental pest because they threaten aquatic biodiversity and conservation, and change ecosystem character and processes. Mature carp are omnivorous and opportunistic feeders. They take whatever food is available, mainly from soft sediments, but also from the water column and the water surface. They

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eat a wide range of food items, mainly benthic and pelagic invertebrates, and have a preference for benthic larvae, such as chironomids, as well as molluscs, large plankton and plants (Sibbing 1991). Carp ingest benthic food items by sucking up mud, then ejecting unwanted particles (Cahn 1929; Sibbing 1991). This action can create crater-like depressions and is effective to 12 cm below the sediment surface (Sibbing 1991). Like all cyprinids, carp lack oral teeth, a stomach and enzymes to break down cellulose, and are illequipped to bite, shred or chew fibrous food such as plant leaves (Sibbing 1991). Nevertheless, physical mastication is an important part of food processing and carp can masticate using specialised structures towards the back of the mouth. They can also crack hard items such as seeds and molluscs. Carp spawn in shallow water over recently flooded vegetation, including wetlands during rising water levels or on floodplains after overbank flooding (Balon 1975). Spawning in carp is therefore linked to flooding, which generally occurs in spring in temperate regions. Spawning can occur only once the critical water temperature of 18oC has been reached (Balon 1975). Carp move to find suitable spawning habitats, covering only short distances if confined to a wetland, or migrating several kilometres up a river system. These spawning migrations offer

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Effects of Biotic Pressures on Wetland Functioning an opportunity for controlling carp abundance through their exclusion by barriers, by harvesting or by allowing certain fish to pass at size-selective gates. Carp eggs are small, and fecundity increases with age, body size and condition. A mature female produces several hundred thousand eggs per season. There are numerous domestic strains of carp, bred for pond culture in different environments. Comparisons of European and Israeli pond strains show that these are genetically distinct from each other, and that the pond strains are distinct from European wild carp (Kohlman 1999), however no comparisons have yet been made to determine genetic similarities of pond and wild carp to feral carp. Knowledge of these domestic genotypes can be used to establish possible origins of feral carp, to interpret dispersal patterns and to identify likely vectors, all of which can assist in formulating policy to restrict their impact. In Australia, for example, feral carp in the Murray-Darling Basin appear to be mainly one of three introduced strains (Davis et al. 1999). Although growth and physiological characteristics of the domestic strains are well known, this knowledge has not been used in relation to feral carp, for example to understand their success, or to assess ecological risks and impact. This can force a conservative approach, as in New Zealand, where feral koi carp are considered as much as a threat as non-koi carp. The present inter-continental distribution of feral carp is due to their deliberate and accidental introduction by humans for a number of reasons (e.g. as an aquarium fish, or for food). The role of human vectors in carp distribution continues today, largely driven, as in historical times, by the actual or potential value or use of the fish. Some movements are legal, such as the international trade in live fish, but some are illegal (depending on a country’s regulations), such as live stocking of fish ponds, transport of trophies, use of live bait and release of live individuals. In contrast to North America and Australia, feral carp are not considered a problem in European wetlands, except locally in parts

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of England and southern France. This may be attributed to differences in abundance. Although only a few biomass data are available for feral carp in Australia and North America, these show that carp can reach more than 400 kg ha−1, which is much higher than in most European rivers and wetlands. Supplementary feeding would be required to maintain such stocking densities in pond aquaculture (Roberts 1998), which is a good indicator that feral carp are placing strong feeding pressures on their new habitats. Reasons for such high abundances in North America and Australia have not been specifically established. Reasons are likely to be specific to the localities, and include characteristics of the species and the environment. Species characteristics that contribute to their high abundance are high fecundity, a bet-hedging reproductive strategy (repeat spawning, which ensures several cohorts within one season), omnivory and wide environmental tolerance. Environmental characteristics which favour carp include fewer or no diseases and predators, and disturbed or altered environments. River regulation is one type of environmental disturbance that appears to favour carp, although the reasons for this have not been comprehensively established (Gehrke et al. 1995). The threat posed by feral carp to wetlands is loss of diversity, loss of structural elements, habitat modification and, arising from all of these, altered ecological processes (Koehn 2004). The impact is due principally to the benthic feeding of mature carp; although juveniles are also benthic feeders, their impact is much less owing to their small size, smaller food items and less powerful feeding action. The direct consequences of carp feeding are depleted benthic invertebrates, particularly molluscs and chironomid larvae, and fewer plant propagules, such as seeds and turions, although the evidence is not always obvious (Fletcher et al. 1985; Roberts 1998; Koehn 2004). This is accompanied by loss of submerged macrophytes, as foraging around plant roots dislodges smaller and less robust plants, particularly seedlings and submerged species, typically

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in the Characeae and Potamogetonaceae. The benthic feeding of carp also leads to increased water column turbidity, the transfer of nutrients or other materials adsorbed to sediment particles from substrate to the water column and a loss of refugia for small fish and large zooplankton. Over the longer term, these impacts can affect the thermal characteristics of wetlands, nutrient cycling, the role of sediments and trophic structure (Roberts 1998). Impact intensity is determined by carp density within the wetland (Lougheed et al. 1998), and higher carp densities mean greater feeding pressure and more intense foraging activity. Field studies of carp impacts on macrophytes have attempted to define a critical density of carp, but this approach is too simplistic, as it takes no account of carp behaviour and how this is altered by season or modified by different environments. Climate is another factor relevant to carp impacts and management. In warm climates, there is no cessation of activity or growth over winter. Continuous growth results in overall faster growth rates, and different age– size relationships; in turn this has forced the development of new data sets and associated techniques, such as using otoliths rather than scales for aging (Vilizzi and Walker 1999). Although density and level of activity determine the intensity of impact from foraging carp in a particular environment, reducing, or even eliminating, carp is not a guarantee that the wetland will self-restore. If carp disturbance has been particularly intense or long-term, then

certain critical self-regulatory thresholds may have been passed. In the case of macrophytes, for example, natural regeneration is unlikely if seeds or other propagules are lost or if the water is too turbid for growth. Instead, active management is required, such as planting out or manipulating the water regime (Roberts 1998). In determining whether to implement a carp control plan, it is important to distinguish symptoms from causes. Although wetlands and shallow lakes dominated by carp typically have low water clarity and lack macrophytes, degraded wetlands without carp can also show these symptoms (Lougheed et al. 1998). It is also necessary to consider the history of other wetland changes as these may be stronger controlling factors. For example, impounding water in an isolated floodplain wetland for agriculture changes the water regime from intermittent to permanent, providing secure growing conditions for carp (Roberts 1998). In this case, it would be more effective to manage the water regime than the carp. Carp can be managed at the level of individual wetlands by altering fish abundance, but successful management rarely involves managing the species in isolation from other environmental impacts or from other carp populations in the watershed. Carp are a symptom as well as a cause of environmental degradation. Reducing their impacts requires an understanding of their role in the larger suite of events that have led to degradation of a particular aquatic environment.

Case study 3 The white-headed duck and the threat from the ruddy duck in Europe Baz Hughes Wildfowl and Wetlands Trust, Slimbridge, United Kingdom The white-headed duck (Oxyura leucocephala) is a globally threatened species (IUCN 2004). Numbers in the central Asian populations have declined markedly since the 1930s, from around 100 000 to less than 10 000 birds (Green and Hunter 1996). The Spanish population fell to

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400 birds in 1950 then to only 22 birds in 1977. Destruction of habitat and excessive hunting were the main causes of decline. Habitat protection, protection from hunting and captive breeding have subsequently increased numbers of the white-headed duck in Spain to over 2500

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countries also hold the greatest numbers of records. Ruddy duck have now been recorded in 21 Western Palearctic countries (Hughes et al. 2006) with breeding records in at least 11, and regular breeding attempts in 6 (France, Ireland, Morocco, the Netherlands, Spain, and the UK). However, outside the UK, only France holds a significant number of breeding pairs (c. 20). Large flocks of wintering birds now occur in Spain and France. In January 1997, about 30 ruddy ducks were recorded in northern Spain following freezing conditions across northern Europe, and up to 200 birds have wintered annually at Lac de Grand-Lieu in northern France since 1995/96 (Hughes et al. 2006). There are three possible sources of the ruddy duck reaching continental Europe: feral birds from the UK; captive birds from Europe; and natural transatlantic vagrants from the USA. DNA fingerprinting has recently proved that the birds in Spain do not originate from North America (V. Muñoz, personal communication) and that all captive and feral ruddy ducks in Europe are probably descended from the seven birds imported into the UK in the 1940s. Although relatively large numbers of ruddy duck are kept in captivity on the continent, most keepers hold few birds and most are feather clipped or pinioned so that they cannot fly. It is therefore highly unlikely that the appearance (and subsequent disappearance) of large flocks of ruddy duck in France and Spain could be explained by mass escapes from captivity. There is supporting evidence to suggest a British source, including

birds (Hughes et al. 2006). Other European conservation initiatives for the white-headed duck have taken place, for example habitat restoration and reintroduction programmes in France and Italy. Ruddy duck (Oxyura jamaicensis) are common and widespread in their native habitat in North America, where there is an increasing population of over half a million (Wetlands International 2002). In the 1940s, Peter Scott imported three pairs of ruddy duck to Slimbridge, UK. Some of their ducklings managed to evade capture for wing-clipping, and escaped. These formed a feral population in the UK, which now numbers some 5000 birds (Hughes et al. 2006). A survey of breeding ruddy duck in the UK in 1994 estimated a total population of about 900 pairs (Hughes et al. 1998). Most birds were found on Anglesey, in the northern counties of England, and in southern Scotland, representing a northward shift in breeding distribution since the BTO Atlas survey of 1988– 1991. Annual spring and summer records of ruddy duck on the Shetland and Orkney Isles and in Iceland during the 1990s suggest this northward range extension may be spreading further afield. In 1965, the first European record of a free-flying ruddy duck outside of the UK was reported from Sweden. As the numbers of ruddy duck in the UK increased, so did the number of sightings abroad (Figure 29.5). This was initially in countries closest to the UK, such as France, the Netherlands, Belgium and Ireland, and these

200

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160 3000 120 2000 80 1000 40 0

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0 1960

Fig. 29.5 Numbers of ruddy duck in the UK in relation to the number of continental records (after Hughes et al. 1999). Open circles – UK population; closed squares – continental records.

UK population

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large winter ruddy duck concentrations overseas, the fact that records in Europe mirror the growth of the UK population (even to the extent that the number of continental records declined in years following declines in the UK ruddy duck population), that the European records are in countries close to the UK, and that there are annual spring and summer records on the Northern Isles and in Iceland, all suggest that ruddy duck in continental Europe originate mainly from the UK feral population. They may be in the early stages of developing the long distance migratory strategy shown by their wild ancestors in North America. The risk to the white-headed duck from ruddy duck has arisen because the two species hybridise readily, with the ruddy duck possessing the competitive edge (Figure 29.6). This threat from the ruddy duck is recognised as the most important threat to the white-headed duck in the Global White-Headed Duck Conservation Action Plan (Hughes et al. 2006). At present, hybridisation between ruddy and white-headed ducks only occurs in the western Mediterranean, and local control of ruddy duck and hybrids in Spain, Portugal, Morocco and France is reducing the threat. This effort is becoming increasingly difficult as ruddy duck numbers in Europe, and in the UK in particular, increase. If ruddy duck enter the Asian white-headed duck population

in any numbers, control would be logistically impossible. DNA analysis has revealed that ruddy duck and white-headed duck are clearly different species and that they have been geographically isolated for 2–5 million years (McCracken et al. 2000). Ruddy duck and white-headed duck are genetically as different from each other as are Baikal Teal Anas formosa and Brazilian Teal Amazonetta braziliensis, the former from the Far East, the latter from South America. Following the first record of ruddy duck in Spain in 1983, small numbers occurred during the 1980s, and then numbers increased to some 20 birds per year in 1991–1993. Relatively few of these were controlled (Figure 29.7), and some interbred with white-headed duck. The number of hybrids recorded thus increased quickly with 31 killed in 1992 and 1993. Subsequently, a mean of 14 ruddy ducks and hybrids have been shot annually (range 4–30). To the end of 2003, at least 122 ruddy ducks and 61 hybrids had been controlled in Spain. The number of countries taking action against ruddy duck has increased significantly in recent years. By 2004, at least 14 countries in the Western Palearctic had taken some action to control the species (Belgium, Denmark, France, Hungary, Iceland, Ireland, Italy, Morocco, the Netherlands, Portugal, Spain, Sweden, Switzerland and the United Kingdom). This compares with only 6 countries in 1999. At least 352 ruddy duck and hybrids have now been controlled in 6 countries 30 Ruddy ducks Hybrids

25 20 15 10

Fig. 29.6 Hybrid ruddy duck × white-headed duck are subject to control programmes in Spain. (Photo by M. Hulme.)

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2003

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Fig. 29.7 Numbers of ruddy duck and ruddy duck × white-headed duck hybrids shot in Spain, 1984–2003.

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Effects of Biotic Pressures on Wetland Functioning excluding the UK (Denmark – 1, France – 160, Iceland – 3, Morocco – 2, Portugal – 3, and Spain – 183) and a further three countries have indicated that attempts will be made to shoot birds if they arrive (Hungary, Italy, Slovenia). Concerted eradication programmes are in operation in four countries (France, Portugal, Spain and the UK) and one is planned in Morocco. Since 1999, 4200 ruddy duck have been shot in the UK, and the government is considering

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whether a national eradication programme should take place. The ruddy duck has now been listed on Annex B of the EC CITES Regulations (338/97) on the grounds that it poses an ecological threat to indigenous species. This gives member states the opportunity to ban or place restrictions on the keeping of ruddy ducks in captive collections. Only with the eradication of both captive and feral populations will the future of the white-headed duck be secure.

Spartina anglica: a young successful species invading coastal salt marshes M.L. Ainouche, A. Baumel Botany Laboratory, University of Rennes, Rennes, France

Spartina anglica C.E. Hubbard (Poaceae) represents a classical example of a young invasive species (one century old) resulting from hybridisation followed by polyploidisation, and now colonising tidal mud-flats and salt marshes in several continents. The origin of this species has been well documented on the basis of morphology (Marchant 1967; Hubbard 1968), chromosome (Marchant 1968) and isozyme data (Guénégou et al. 1988; Gray et al. 1990; Raybould et al. 1991a,b). It arose around 1890 in southern England (Southampton bay), following hybridisation between the indigenous Spartina maritima Fernald (2n = 62) and the introduced east North American Spartina alterniflora Loiseleur (2n = 60), the latter being the maternal genome donor (Ferris et al. 1997). Chromosome doubling of the subsequent hybrid, Spartina × townsendii, gave rise to the new double genome (allopolyploid) species S. anglica (2n = 120, 122 and 124, according to Marchant (1968)), a fertile, vigorous and variable-form clonal plant (Thompson 1991) colonising British salt marshes and estuaries since 1890. The individual plants may be very variable in size, with numerous rhizomes forming a robust underground system producing culms. S. anglica bears longer and broader leaves, standing at a wider angle with the culm, than in its parent S. maritima.

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Spartina anglica rapidly became a common pioneer of British low-lying intertidal mudflats. It tolerates a wide range of conditions across the successional sequence of salt marsh zones. The species can tolerate regular submersion in salt water for up to 6 hours per tide, which enables it to extend much further down the shore than other plants, and therefore to create an extended niche (Thompson 1991). Plants also grow easily in brackish water, forming vigorous populations in estuaries. The high productivity of this species, combined with the rapid growth of its rhizomes, led to the formation of an extensive network that accretes large volumes of sediments. Since its formation, S. anglica has spread along the coast of the British Isles, whereas both parent species (S. alterniflora and S. maritima) remain confined to few sites along the British southern coast (Gray and Raybould 1997). The wider extremes of habitat (at local and geographical scales) tolerated by the hybrid, compared to its native parent species, indicates higher invasive potential. S. anglica has now a worldwide distribution, as it spreads by both natural means and deliberate introductions by humans (north Europe, Australia, New Zealand, China and others) for land reclamation. This has resulted in various local policies in order to manage the species, including attempts at chemical or physical eradication.

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Much debate has occurred on both the reasons for the success of this species and its long-term evolutionary potential (Gray et al. 1997; Raybould et al. 1998). Amounts of genetic diversity available to a new species determine the potential for its subsequent evolution. Molecular investigations have revealed that this species lacks genetic variation at both coding (Raybould et al. 1991a) and non-coding portions of the genome (Ainouche et al. 1999; Baumel et al. 2001), reflecting a severe genetic bottleneck at the time of the species formation. Most of the populations so far examined in the native region of the species (Western Europe) display the same pattern

of DNA (Baumel et al. 2001). This genotype, even growing in pioneer situations, will have to face biotic as well as abiotic pressures, such as in Europe where populations may be largely infected by ergot fungus, Claviceps purpurea (Raybould et al. 1998). Spartina anglica represents a system where genetic diversity is restricted to the (intra-individual) subgenomic level (i.e. within its own DNA), as the hybrid genome contains two different duplicated subgenomes from S. alterniflora and S. maritima. The future of this species will be highly dependent on the long-term evolutionary dynamics of its hybrid genome, at both the structural and expression levels.

Case study 5 Salvinia molesta: the wetland space invader David Mitchell Charles Sturt University, Albury, Australia The surface-floating aquatic fern, Salvinia molesta D.S. Mitch. is native but not common in South America (Forno and Harley 1979; Forno 1983). However, over the last half of the twentieth century it has progressively invaded tropical and subtropical waters in many countries in Africa, South East Asia, the Asian Pacific and Australasia, and recently in the United States of America. It is now a major aquatic weed rivalling water hyacinth, even though there are effective control measures for most situations. The plant was placed by Herzog (1935) in the species S. auriculata Aubl., which he distinguished from other species in the genus by the presence of the inverted egg-beater like arrangement of four hairs at the tip of papillae on the upper surface of the leaf. This is still a primary distinguishing feature of the species, as the other species in the polyploid S. auriculata complex have not spread adventively from their native distribution and only S. auriculata is widespread and common through South America and into Mexico (Mitchell and Thomas 1972; Forno 1983). Loyal and Grewal (1966), working

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with Indian material, showed the taxon to be a sterile pentaploid, which maintains itself by vegetative apomixis and, in 1972, the plant was recognised as a separate species on the basis of the sterility of the sporocarps and their branching pattern (Mitchell 1972). It is noteworthy that only one species in the genus, S. molesta, has exhibited aggressive invasive behaviour in alien situations, as in many respects the species are very similar to one another. Mitchell and Tur (1975) showed that the plant was capable of doubling its biomass in 3.4 days in laboratory cultures and in 8–12 days in field conditions (Lake Kariba, central Africa) and these rapid rates of growth are likely to be part of the reason for its success as an invading species (Figure 29.8), especially when assisted by a strategy of effective vegetative reproduction involving repetitive plant breakage at the nodes (Room 1983). While S. molesta responds with rapid growth in nutrient rich conditions, it is also capable of surviving for many months in the low nutrient conditions typical of many tropical waters.

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Effects of Biotic Pressures on Wetland Functioning

Fig. 29.8 Salvinia (Salvinia molesta) has spread from its native range in South America to many tropical and subtropical wetlands and lakes where it can completely cover open water areas, reducing light penetration, resulting in deoxygenation and blocking transport. (Photo by author.)

In addition, dead plant tissue provides very effective insulation against heat and drying, enabling dormant buds to survive such adverse conditions for long periods. The main, and probably only, effective dispersal agent from one water body to another is human. The plant was initially spread as an aquarium plant, garden plant and botanical curiosity, but now most movements from one water body to another occur inadvertently on boats and similar equipment. It spreads within a water body by wind and water current, and growth is independent of depth. In still conditions it will rapidly colonise areas of open water, where it can provide a base for the growth of a number of emergent aquatic plants leading to the formation of floating ‘sudd’ islands (Boughey 1963). Salvinia molesta is a serious environmental and economic weed. Its capacity to occupy the whole surface of smaller water bodies and extensive areas of larger water bodies drastically alters the aquatic habitat for many of its normal occupants, by shutting out light and causing low oxygen conditions in the water column, as outlined, for example, for Kakadu National Park, northern Australia, by Storrs and Julien (1996). Similarly, large mats of the weed interfere with navigation, fishing and

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tourism and can adversely affect water quality for human consumption. The most profound social consequences of the invasion of a water body occurred in the Sepik River, Papua New Guinea, in the 1970s and 1980s (Gewertz 1983). The native inhabitants of the regions bordering the river system and its associated lakes were almost totally dependent on these systems for food and access to their villages. Unfortunately, the dugout canoes they used for transport could not move through the dense mats that surrounded their villages, leading to their partial or complete abandonment and occasioning great emotional distress. Fortunately, effective control measures are now available for management of nuisance infestations of Salvinia molesta in most situations. Biological control by a weevil, Cyrtobagous salviniae has achieved spectacular success, for example, in the Sepik River system in Papua New Guinea, where 250 km2 was reduced to less than 2 km2 in about 18 months (Room 1986). In Kakadu National Park, the extent of the weed and the balance between it and the control agent is affected by the timing of the onset of the wet season and of the main flooding events. In some years, the weevil population crashes early, allowing the Salvinia to recover and extend rapidly before the weevil population recovers sufficiently to re-establish control. In other years a good balance is maintained and little excessive growth of the weed occurs (Storrs and Julien 1996). In more temperate conditions, as near Moruya, 35°55´ South, on the southeast coast of Australia, the weevil population increases too slowly in the summer and does not survive the winter. The most effective chemical control is achieved by a formulation that mixes kerosene with an anionic surfactant (Kemmat) at 50 : 1. This burns and sinks the weed. An addition of a herbicide, such as diuron (3-(3,4-dichlorophenyl)-1,1-dimethylurea) applied at rates of 0.1–0.25 kg ha−1 kills the plant (Diatloff et al. 1979). Physical control methods are ineffective as they break the plant, but preventive measures directed at minimising

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the spread of the plant by publicising its dangers can also assist. Strategies that integrate all these methods, as is most appropriate for particular circumstances, can be very effective, providing they are rigorously and persistently applied. The rapid spread of Salvinia molesta in the last six decades and its recent invasion of the

CR IT ICAL L E S S ON S There are many lessons that can be learnt from investigating past biotic disruption of wetlands by invasive species. A few of those that are considered most critical are listed below: • Prevention and early intervention are more likely to be successful or cheaper than later control and management. In this respect special attention should be directed towards areas that are disturbed by other activities. • Coordinated and multi-sectoral approaches that incorporate effective legislative, educational and coordination steps are more likely to result in partnerships and actions for effective management than are individualistic approaches. • Management strategies and actions require both a concerted and ongoing effort and should be based on a sound information base with the latter being compiled by ongoing and formalised risk assessment and ecological investigations. These need to consider that wetland dysfunction can be expressed by adverse change at the species and habitat level and also through disruption of wetland functions and processes. • Control measures are generally location and species specific, and local information and experience can prove invaluable, especially if backed by adequate information and management resources. • Whilst a single control measure may prove successful, there is often a need for an integrated programme that incorporates different techniques and sectors. An over-reliance on one method, whether it be chemical, biological or physical, could prove disastrous.

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southern states of the USA demonstrate the potential dangers it presents to tropical and subtropical surface waters throughout the world. Even though it can be effectively controlled there is no room for complacency, as control is expensive and very rarely results in complete eradication of the weed.

• A careful cost–benefit analysis should precede all control programmes and be based on the best available ecological knowledge and understanding of management options. It should consider the need for rehabilitation of areas where control has been successful. • Balanced against a need for early intervention and prevention is the wisdom of deciding whether any action is needed or is likely to be successful. Such decisions need to consider not only ecological and economic factors but also the social and psychological factors. • Managers should consider both the non-ecological and ecological causes of biotic pressures, and direct resources to address long-term strategic outcomes as well as tactical short-term outcomes.

R EFER EN CES Ainouche M.L., Baumel A. and Bayer R.J. 1999. Molecular investigations in the young allopolyploid species Spartina anglica Hubbard (Poaceae) in France. Abstracts 4028, XVI International Botanical Congress. Saint Louis, Missouri (USA), 1–7 August 1999, available at http://www.biologie.uni-hamburg.de/ b-online/ibc99/ibc/abstracts/listen/abstracts/4028. html, last accessed on 22 January 2009. Balon E.K. 1975. Reproductive guilds of fishes: a proposal and definition. Journal Fisheries Research Board, Canada 32, 821–864. Barrett S.C.H. and Richardson B.J. 1986. Genetic attributes of invading species. In: Groves R.H. and Burdon J.J. (editors), Ecology of Biological invasions: An Australian Perspective. Australian Academy of Science, Canberra, pp. 21–33.

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Effects of Biotic Pressures on Wetland Functioning Baumel A., Ainouche M.L. and Levasseur J.E. 2001. Molecular investigations in populations of Spartina anglica C.E. Hubbard (Poaceae) invading coastal Brittany (France). Molecular Ecology 10(7), 1689–1701. Beringer J.E. 2000. Releasing genetically modified organisms: will any harm outweigh any advantage? Journal of Applied Ecology 37, 207–214. Boughey A.S. 1963. The explosive development of a floating weed vegetation on Lake Kariba. Adansonia 3, 49–61. Cahn A. 1929. The effect of carp on a small lake: the carp as a dominant. Ecology 10, 271–275. Clout M.N. and Lowe S.J. 2000. Invasive species and environmental changes in New Zealand. In: Mooney H.A. and Hobbs R.J. (editors), Invasive Species in a Changing World. Island Press, Washington, DC, pp. 369–383. Cook D.D., Setterfield S.A. and Maddison J.P. 1996. Shrub invasion of a tropical wetland: implications for weed management. Ecological Applications 6, 531–537. Cowie I.D. 1996. Weed ecology. In: Finlayson C.M. and Von Oertzen I. (editors), Landscape and Vegetation Ecology of the Kakadu Region, Northern Australia. Geobotany 23, Kluwer Academic Publishers, Dordrecht, pp. 113–135. Davis K.M., Dixon P.I. and Harris J.H. 1999. Allozyme and mitochondrial DNA analysis of carp Cyprinus carpio L., from south-eastern Australia. Marine and Freshwater Research 50, 253–260. Delany S., Reyes C., Hubert E., Pihl S., Rees E., Haanstra L. and van Strien A. 1999. Results from the International Waterbird Census in the Western Palearctic and Southwest Asia 1995 and 1996. Wetlands International Publication No. 54. Wageningen, The Netherlands. Diatloff G., Lee A.N. and Anderson T.M. 1979. A new approach for Salvinia control. Journal of Aquatic Plant Management 17, 24–27. Eliot I., Finlayson C.M. and Waterman P. 1999. Predicted climate change sea-level rise and wetland management in the Australian wet-dry tropics. Wetlands Ecology and Management 7, 63–81. Ferris C., King R.A. and Gray A.J. 1997. Molecular evidence for the maternal parentage in the hybrid origin of Spartina anglica C.E. Hubbard. Molecular Ecology 6, 185–187. Finlayson C.M. 1999. Coastal wetlands and climate change: the role of governance and science. Aquatic Conservation: Marine and Freshwater Ecosystems 9, 621–626.

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Finlayson C.M. and Mitchell D.S. 1982. Management of Salvinia molesta in Australia. Australian Weeds 2, 71–76. Finlayson C.M. and Rea N. 1999. Reasons for the loss and degradation of Australian wetlands. Wetlands Ecology and Management 7, 1–11. Finlayson C.M., Julien M.H., Russell-Smith J. and Storrs M.J. 1994. Summary of a workshop on Salvinia molesta in Kakadu National Park, Northern Territory, Australia. Plant Protection Quarterly 9, 114–116. Fletcher A.R., Morison A.K. and Hume D.J. 1985. Effects of carp (Cyprinus carpio L.) on aquatic vegetation and turbidity of waterbodies in the lower Goulburn River Basin. Australian Journal of Marine and Freshwater Research 36, 311–327. Forno I.W. 1983. Native distribution of the Salvinia auriculata complex and keys to species identification. Aquatic Botany 17, 71–83. Forno I.W. and Harley K.L.S. 1979. The occurrence of Salvinia molesta in Brazil. Aquatic Botany 6, 185–187. Gehrke P.C., Brown P., Schiller C.B., Moffatt D.B. and Bruce A.M. 1995. River Regulation and Fish Communities in The Murray-Darling River System, Australia. Regulated Rivers: Research and Management 11, 363–375. Gewertz D.B. 1983. Sepik River Societies. Yale University Press, New Haven. Gitay H., Suarez A., Watson R., Ansimov O., Chapin F.S., Victor Cruz R., Finlayson M., et al. 2002. Climate Change and Biodiversity. IPCC Technical Paper V, Intergovernmental Panel on Climate Change, Geneva, Switzerland. Goldschmidt T., Witte F. and Wanink J. 1993. Cascading effects of the introduced Nile perch on the detritivorous/phytoplanktivorous species in the sublittoral areas of Lake Victoria. Conservation Biology 7, 686–700. Gopal B. 1987. Water Hyacinth. Elsevier, Amsterdam. Gray A.J., Benham P.E.M. and Raybould A.F. 1990. Spartina anglica – the evolutionary and ecological background. In: Gray A.J. (editor), Spartina anglica – a Research Review. Institute of Terrestrial Ecology, Natural Environment Research Council, Swindon, UK, pp. 77–79. Gray A.J. and Raybould A.F. 1997. The history and evolution of Spartina anglica in the British Isles. In: Patten K. (editor), Proceedings of the Second International Spartina Conference. Washington State University, Cooperative Extension, Olympia, WA, pp. 13–16.

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Lougheed V.L., Crosbie B. and Chow-Fraser P. 1998. Predictions on the effect of common carp (Cyprinus carpio) exclusion on water quality, zooplankton, and submergent macrophytes in a Great Lakes wetland. Canadian Journal Fisheries and Aquatic Sciences 55, 1189–1197. Lowe-McConnel R.H. 1987. Ecological Studies in Tropical Fish Communities. Cambridge University Press, Cambridge, 382 pp. Loyal D.S. and Grewal R.K. 1966. Cytological study on sterility in Salvinia auriculata Aublet, with a bearing on its reproductive mechanism. Cytologia 31, 330–338. Manchester S.J. and Bullock J.M. 2000. The impacts of non-native species on UK biodiveristy and the effectiveness of control. Journal of Applied Ecology 37, 845–864. Marchant C.J. 1967. Corrected chromosomes numbers for Spartina × townsendii and its parent species. Nature 31, 929. Marchant C.J. 1968. Evolution in Spartina (Gramineae). II. Chromosomes, basic relationships and the problem of Spartina × townsendii agg. Journal of the Linnaean Society 60, 381–409. Matthews S. 2004. Tropical Asia Invaded: The Growing Danger of Alien Invasive Species. The Global Invasive Species Programme, Cape Town, South Africa. Mathhew S. and Brand K. 2004. Africa Invaded: The Growing Danger of Alien Invasive Species. The Global Invasive Species Programme, Cape Town, South Africa. McCracken K.G., Harshman J., Sorenson M.D. and Johnson K.P. 2000. Are Ruddy Ducks and Whiteheaded Ducks the same species? British Birds 93, 396–398. McNeely J.A., Mooney H.A., Neville L.E., Schei P. and Waage J.K. (editors) 2001. A Global Strategy on Invasive Alien Species. IUCN Gland, Switzerland, and Cambridge, UK. Mitchell D.S. 1972. The Kariba weed: Salvinia molesta. British Fern Gazette 10, 251–252. Mitchell D.S. 1978. Aquatic Weeds in Australian Inland Waters. Australian Government Printing Service, Canberra. Mitchell D.S. and Thomas P.A. 1972. Ecology of water weeds in the neotropics, International Hydrological Decade Technical Papers in Hydrology 12. UNESCO, Paris, 50 pp. Mitchell D.S. and Tur N.M. 1975. The rate of growth of Salvinia molesta (S. auriculata auct.) in laboratory and natural conditions. Journal of Applied Ecology 12, 213–225.

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Effects of Biotic Pressures on Wetland Functioning Mooney H. 2001. Invasive alien species – the nature of the problem. In: Secretariat of the Convention on Biological Diversity (editors), Assessment and Management of Alien Species That Threaten Ecosystems, Habitats and Species, pp. 1–2. Abstracts of keynote addresses and posters presented at the sixth meeting of the Subsidiary Body on Scientific, Technical and Technological Advice, held in Montreal, Canada, from 12 to 16 March 2001. Montreal, SCBD. CBD Technical Paper No. 1. Newsome A.E. and Noble I.R. 1986. Ecological and physiological characters of invading species. In: Groves R.H. and Burdon J.J. (editors), Ecology of Biological Invasions: an Australian Perspective. Australian Academy of Science, Canberra, pp. 1–20. Ogutu-Ohwayo R. 1999. Nile perch in Lake Victoria: balancing the costs and benefits of aliens. In: Sandlund O.T., Sche P.J. and Viken A. (editors), Invasive Species and Biodiversity Management, pp. 47–63. Kluwer, Dordrecht. Pitttock J., De Poorter M. and Finlayson M. 1999. Workshop on mitigating the impact of alien/invasive species. In: Report of the Thirteenth Global Biodiversity Forum, San Jose, Costa Rica, May 1999, pp. 15–22. IUCN-The World Conservation Union, San Jose, Costa Rica and Gland, Switzerland, available at http://books.google.es/books?id=37Eo5k858EM, last accessed on 22 Jan 2009. Ramsar 1999. Resolution VII.14 on Invasive Species and Wetlands. In: Seventh Meeting of the Conference of the Contracting Parties to the Convention on Wetlands. San José, Costa Rica, 10–18 May 1999, available at www.ramsar.org/res/key_res_vii.14e. htm, last accessed on 22 January 2009. Raybould A.F., Gray A.J., Lawrence M.J. and Marshall D.F. 1991a. The evolution of Spartina anglica C.E. Hubbard (Gramineae): origin and genetic variability. Biological Journal of the Linnaean Society 43, 111–126. Raybould A.F., Gray A.J., Lawrence M.J. and Marshall D.F. 1991b. The evolution of Spartina anglica C.E. Hubbard (Gramineae): genetic variation and status of the parental species in Britain. Biological Journal of the Linnaean Society 44, 369–380. Raybould A.F., Gray A.J. and Clarke R.T. 1998. The long term epidemic of Claviceps purpurea on Spartina anglica in Poole Harbour: pattern of infection, effects on seed production and the role of Fusarium hetrosporum. New Phytologist 138, 497–505. Rea N. and Storrs M.J. 1999. Weed invasions in wetlands in Australia’s Top End: reasons and solutions. Wetlands Ecology and Management 7, 47–62.

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Ridgill S.C. and Fox A.D. 1990. Cold Weather Movements of Waterfowl in Western Europe. IWRB Special Publication No. 13, Slimbridge, UK. Roberts J. 1998. Carp: sifting through the issues. In: Williams W.D. (editor), Wetlands in a Dry Landscape: Understanding for Management. Environment Australia, Biodiversity Group, Canberra, Australia, pp. 253–258. Room P.M. 1983. Falling apart as a life style – the rhizome architecture and population growth of Salvinia molesta. Journal of Applied Ecology 17, 349–365. Room P.M. 1986. Biological control is solving the world’s Salvinia molesta problems. In: Proceedings European Weed Research Society 7th International Symposium on Aquatic Weeds. European Weed Research Society, Loughborough, UK, pp. 271–276. RSPB 1997. Workshop on the Impacts of Climate Change on Flora and Fauna. Royal Society for the Protection of Birds, Sandy, UK. Sculthorpe C.D. 1967. The Biology of Aquatic Vascular Plants. Edward Arnold, London. Sibbing F.A. 1991. Food capture and oral processing. In: Winfield I.J. and Nelson J.S. (editors), Cyprinid Fishes: Systematics, Biology and Exploitation. Fish and Fisheries Series 3, Chapman and Hall, London, pp. 377–412. Storrs M.J. and Julien M.H. 1996. Salvinia. A handbook for the integrated control of Salvinia molesta in Kakadu National Park. Northern Landscapes Occasional Papers, Paper No. 1. Australian Nature Conservation Agency, Darwin. Thompson J.D. 1991. The biology of an invasive plant. What makes Spartina anglica so successful? Bioscience 41, 393–401. van Dam R.A., Finlayson C.M. and Humphrey C.L. 1999. Wetland risk assessment: a framework and methods for predicting and assessing change in ecological character. In: Finlayson C.M. and Spiers A.G. (editors), Techniques for Enhanced Wetland Inventory, Assessment and Monitoring, pp. 83–118. Supervising Scientist Report 147, Supervising Scientist Group, Canberra. van Dam R., Gitay H., Finlayson M., Davidson N.J. and Orlando B. 2002a. Climate Change and Wetlands: Impacts, Adaptation and Mitigation. Background document (DOC.SC26/COP8-4) prepared by for the 26th Meeting of the Standing Committee of the Convention on Wetlands (Ramsar Convention) held in Gland, Switzerland, 3–7 December 2002, available at http://www.ramsar.org/cop8_doc_11_e.htm, last accessed on 25 January 2009.

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van Dam R.A., Walden D.J. and Begg G.W. 2002b. A Preliminary Risk Assessment of Cane Toads in Kakadu National Park. Supervising Scientist Report 164, Supervising Scientist, Darwin, Australia. Vilizzi L. and Walker K.F. 1999. Age and growth of the common carp, Cyprinus carpio, in the River Murray, Australia: validation, consistency of age interpretation, and growth models. Environmental Biology of Fishes 54, 77–106. Waage J. 2001. Outputs of GISP Phase I and future plans of the Global Invasive Species Programme. In: Secretariat of the Convention on Biological Diversity (editors), Assessment and Management of Alien Species that Threaten Ecosystems, Habitats and Species, pp. 3–6. Abstracts of keynote addresses and posters presented at the sixth meeting of the Subsidiary Body on Scientific, Technical and Technological Advice, held in Montreal, Canada, from 12 to 16 March 2001. Montreal, SCBD. CBD Technical Paper no. 1.

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Walden D., van Dam R., Finlayson M., Storrs M., Lowry J. and Kriticos D. 2004. A Risk Assessment of the Tropical Wetland Weed Mimosa pigra in Northern Australia. Supervising Scientist Report 177, Supervising Scientist, Darwin, Australia. Wetlands International. 2002. Waterbird Population Estimates – (3rd edition). Wetlands International, Global Series No. 12, Wageningen. Williamson M. 1996. Biological Invasions. Chapman and Hall, London, UK. Witte F., Goldschmidt T. and Wanink J.H. 1995. Dynamics of the haplochomine cichlid fauna and other ecological changes in the Mwanza Gulf of Lake Victoria. In: Pitcher T.J. and Hart P.J.B. (editors), The Impact of Species Changes in African Lakes. Chapman and Hall, London, pp. 83–110. Wittenburg R. and Cock M.J.W. (editors) 2001. Invasive Alien Species: a Toolkit of Best Prevention and Management Practices. CAB International, Wallingford.

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30 Human Impacts: Farming, Fire, Forestry and Fuel HANS J O O STEN Institute of Botany and Landscape Ecology, University of Greifswald, Germany

IN T R O D U CT ION The intense relationship between humans and wetlands is as old as humanity itself. The most important fossils of ancestral hominids and early humans have been found at former wetland sites, such as Lake Turkana (Kenya), Olduvai Gorge (Tanzania) and the Hadar region of Ethiopia. The 1.5 million year old Turkana Boy, the most complete skeleton ever found of Homo erectus, was excavated in what had been quiet, shallow water: a lagoon near the edge of a lake or an oxbow of a river. Bones of wetland animals on the find scenes indicate hunting and fishing and provide early evidence of human impact on wetlands (Coles 1990; Leakey and Lewin 1992). Impact of modern humans (Homo sapiens sapiens) dates back to the transition of the last glacial to the present interglacial, some 10 000 (14C) years BP, when the combination of changing climates and expanding and migrating human populations exterminated a considerable part of the global megafauna, including wetland species like the giant beaver Castoroides ohioensis (Martin and Wright 1967). Bog bodies, tools, ornaments, weapons and other archaeological remains found in abundance in wetland sediments and peats testify to the long and intimate relationship between people and wetlands during the whole Holocene (Moore 1987; Coles and Coles 1989; Müller-Wille 1999).

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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This relationship was not unambiguous: wetlands were seen both as life-bringing and life-taking, and as inviting and repelling. In early seventeenth century England, fens were described as: ‘The air nebulous, grosse and full of rotten harres; the water putred and muddy, yea, full of loathsome vermine; the earth spuing, unfast and boggy ...’. But other voices of that time recognised their value in providing fodder for horses, cattle, and sheep, as store of ‘osier, reed and sedge’ and as ‘nurseries and seminaries’ of fish and fowl, from which thousands of people gained their livelihood (Wheeler 1896; Pursglove 1988). Historical accounts describe people who lived in and depended almost entirely on wetlands, from the ‘half amphibious’ Fen Slodgers in the English Fenlands (Wheeler 1896) to the wetland peoples of recent times, such as the Marsh-Arabs (Madan) of Southern Iraq (Thesinger 1964) and the Kolepom people in Irian Jaya (Serpenti 1977). Large scale wetland modification for agriculture started with the origin of rice cultivation in China about 6000 BC (Glover and Higham 1996). The Minyans drained and subsequently cultivated the Kopais basin in Greece 3500 years ago (Knauss et al. 1984). Some centuries later, the Babylonians established municipal reed beds and harvested bulrushes for construction purposes (Boulé 1994). Humans have converted wetlands into drylands to win areas for intensive agriculture and forestry, and to fight water-related parasitic diseases (Appleton et al. 1995). Wetland destruction may result not only from impacts on site, but also from changing land

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use outside the wetland boundaries. Already, 3000 years ago, anthropogenic upland deforestation led to more flashy river discharges of the Rivers Rijn, Maas and Schelde. The consequent widening of the estuaries gave opportunity for the sea to penetrate, and eventually led to massive erosion of the peatlands in the perimarine floodplain of the Netherlands (Pons 1992). Human activities, however, also led to the origin and expansion of wetlands. Many peatlands in Europe came into being because of human interference with landscape hydrology (Moore 1975, 1993; Maltby and Caseldine 1982; Törnqvist and Joosten 1988; Bennett et al. 1998). The first textbook devoted to fish breeding in artificial ponds appeared in 473 BC in China, where fish ponds now cover an area of about 40 000 km2 (Mate˘na and Berka 1987; Opuszyn´ski 1987). The Oostvaardersplassen in the Netherlands (Vera 1988) arose from a whole succession of human activities. The area came into being after the partial pumping of Zuidelijk Flevoland, the last polder built in the Dutch IJsselmeer (1968), a shallow lake that had originated by damming and subsequent sweetening of the former Zuiderzee (1932), a marine area itself formed in the thirteenth century when human induced erosion and floods connected the former inland Lake Flevo to the North Sea via the Wadden Sea. Human impact on wetlands has been varied (cf. Maltby 1986; Gosselink and Maltby 1990; Williams 1990; Dugan 1993; Maltby et al. 1994; Spiers 1999). Table 30.1 presents an overview of these direct and indirect impacts according to four basic types of ecosystem disturbance (Van Leeuwen 1981): • too little input (‘malnutrition’); • too little output (‘constipation’); • too much input (‘pollution’); • too much output (‘deprivation’). Normally, an act principally affects one compartment, but as these are strongly interrelated, the effects commonly spread to other compartments. In Table 30.1 the compartments are arranged hierarchically: changes in higher compartments generally lead to larger effects in the lower compartments than the reverse (Klijn 1995).

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The current extent of wetlands in the world is not well detailed. Of 206 countries or territories recently studied, only 7% had adequate or good inventory coverage (Finlayson and Spiers 1999). Even less is known about the quantitative losses and the current condition of the global wetland resource (cf. Spiers 1999). The OECD (1996) estimates that the world may have lost 50% of its wetlands since 1900. Much of this loss occurred in the northern temperate zone during the first half of the twentieth century. Since the 1950s, tropical and subtropical wetlands, particularly swamp forests and mangroves, have increasingly been lost. More than 80% of the internationally protected Ramsar-wetlands have undergone, or are threatened by, human-induced ecological change (Finlayson and Spiers 1999). With special reference to mires and peatlands, this chapter presents some key aspects of agriculture, forestry and peat extraction as examples of wetland destruction across the globe. Mires have long remained the last wildernesses in many areas of the world. Their limited accessibility (‘too wet to drive and too dry to swim’) protected them against human intervention. They constitute the larger part of global freshwater wetlands (Maltby 1986; Mitch et al. 1994; Lappalainen 1996) and can be regarded as the final terrestrial frontier (Joosten 1999). It is estimated that perhaps 80% of the original resource is still in a largely pristine condition (Safford and Maltby 1998; Joosten 1999), although the tropical resource is rapidly diminishing (cf. Rieley 1999b). In the non-tropical world, Europe is the continent with the largest mire losses, because of its long history, high population pressure and climatic suitability for agriculture. More than 50% of its former 0.6 × 106 km2 of mires is no longer peat accumulating. Since peat extraction, subsidence, oxidation and erosion following human activities have changed many former peatlands into mineral soils (cf. Dömsödi 1988; Leenders 1989; Joosten 1994), possibly 10–20% of the original European mire resource does not exist any more, even as peatlands (Joosten 1999; Joosten and Clarke 2002).

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Incoming sunlight (increasing albedo by cloud formation ⇒ climate change)

Precipitation (climate change) Surface and ground water (by upstream damming, diversion, or extraction from catchment) Tidal dynamics (damming) Flow rate (damming, water extraction) Alkaline water (⇒ acidification) Fresh water (⇒ salinisation) Salt water (⇒ reversed salinisation)

On (macro- and meso-) climate

On bedrock, substrate, and relief

On water level

On water dynamics

On water quality

Atmo-sphere

Geo-sphere

Hydro-sphere

Sediment (decreased supply by upstream damming) Peat (drainage ⇒ accumulation stop)

By too little input of

Compartments Human ⇒ ⇓ impact ⇓ By too much input of

By too much output of

Salts (increased agricultural production and evapotranspiration⇒ decreased runoff ⇒ salinisation)

Wave energy (swamp and marsh destruction ⇒ coastal erosion)

Surface water (downstream damming ⇒ inundation)

Sediment (decreased water flow rates ⇒ siltation)

Water level fluctuations and high water levels (water level control ⇒ less morphogenesis)

Evapotranspiration (climate change) Surface water (drainage, extraction from wetland, canalisation)

Minerals, sediments, peat (mining, dredging, erosion) Height (subsidence) Relief (levelling) Sinuosity (stream canalisation)

(Continued)

Fertilising, acidifying, oxygen Alkaline water (drainage consuming, salinising, ⇒ acidification) surfacing (oil), and toxic Fresh water (extraction ⇒ compounds salinisation) Heath (effluents, cooling water)

Flow pulses and flow rate (increased surface runoff, weir management)

Surface water (increased runoff by alteration of surfaces and deforestation) Sea water (sea level rise by climate change)

Refuse (land fill) Spoil (dredging) Surfacing material (urbanisation) Sediment (upland erosion, mining ⇒ siltation)

Outgoing warmth Greenhouse gases Outgoing sunlight (greenhouse gases and condensation particles (⇒ cloud (increasing albedo cloud formation ⇒ climate formation) by desertification ⇒ change ) climate change)

By too little output of

Table 30.1 Types of human impact, that may substantially change a wetland’s character; with examples of (causes) and (⇒ effects).

Human Impacts 691

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Organic material/ peat (drainage ⇒ disappearance of peat accumulating plants)

On soil conditions

On plants and animals

On humans

Pedo-sphere

Bio-sphere

Noo-sphere

Traditional management (socio-economic changes ⇒ abandonment)

Genetic material (obstruction of migration and diaspore transport) Disturbance (⇒ changes in patch dynamics)

By too little input of

Compartments Human ⇒ ⇓ impact ⇓

Table 30.1 Continued.

People (hygiene, green revolution, mosquito and bilharzia control ⇒ overexploitation)

Plant biomass (omission of mowing or grazing ⇒ colonisation by forbs and trees) Animal biomass (tsetse control ⇒ overgrazing and erosion)

Salts (less water through flux ⇒ salinisation)

By too little output of

People (colonisation, transmigration) ‘Western civilisation’ (assimilation ⇒ loss of indigenous wetland cultures)

Radioactive and UV radiation (ozone hole, ⇒ mutations) Fertilising and toxic compounds Genetic material and biomass (exotic and invasive species) Disturbance (tourism)

Heath (⇒ permafrost thawing)

By too much input of

People (emigration / urbanisation, introduced diseases, genocide ⇒ loss of indigenous wetland cultures)

Biomass (overexploitation) Genetic material (extirpation ⇒ extinction)

Topsoil (erosion, mining) Peatland acrotelm (drainage, fire) Soil horizons (ploughing)

By too much output of

692 HANS JOOSTEN

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Human Impacts Table 30.2 Causes of anthropogenic mire losses in the non-tropical world. (After Joosten 1999.) 103 km2 Agriculture Forestry Peat extraction Urbanisation Inundation Indirect losses (e.g. erosion, landfill) Total

250 150 50 20 15 5 490

% 50 30 10 5 3 1 100

AGR ICU L T UR E Wetland destruction for agriculture Agriculture is still the principal cause of wetland loss worldwide (Table 30.2). By 1985 it was estimated that 56–65% of the available wetland in Europe and North America had been drained for intensive agriculture, alongside 27% in Asia, 6% in South America and 2% in Africa, a total of 26% worldwide (OECD 1996). Wetland destruction for agriculture has been occurring for thousands of years. We learn from Pliny the Elder, that in the year 593 BC the Roman senate ordered consul Cornelius Cethegus to drain wetlands in Latium, about 40 miles from Rome, with help from the army. The legions were employed in these useful labours to preserve the soldiers from the dangerous temptations of idleness. Hannibal (246–182 BC) instructed his troops along the coast of Africa to convert wetlands near Sirmium to plantations of olive trees. The Roman Emperor Claudius (10 BC–54 AD) employed 30 000 men for eleven years in draining Fucino Lake in Italy (Wheeler 1896). The Romans had gained this expertise from the Greeks and the Etruscans and were quick to export it to their colonies. The Emperor Hadrian (76–138 AD) built the Car dyke, a drain encircling the western edge of the English Fens. On the Medway estuary in Kent, banks built by the Romans to keep out the sea lasted until the eighteenth century (Purseglove 1988). In the Netherlands, the first large-scale drainage systems

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693

in Flanders and Zeeland were constructed by the Romans (Borger 1992). Simultaneously, the Mayans developed raised bed agriculture techniques in wetlands (Boulé 1994). Starting from the eighth century, the extensive mires of Holland were colonised and used as meadow, pasture and arable land. As early as 1100 AD the technique of peatland reclamation and colonisation had achieved such success that Dutch expertise was exported to Germany (Borger 1992). An important driving force for reclamation of the marshes and peatlands in various parts of Europe during the Middle Ages was the church, eager to derive economic benefit from this agricultural development (Purseglove 1988). In many places, drainage gradually started to show negative side-effects. River mouths silted because of reduced tidal scouring. The reclamation-induced peat oxidation and subsidence led to a lowering of the peatland surface (Segeberg 1960; Hutchinson 1980; Eggelsmann 1990). This necessitated continuous deepening of ditches and canals (the ‘vicious circle of peatland utilisation’, Kuntze 1982) and caused deteriorated drainage conditions and increased flooding intensity. People were forced to take defensive measures by constructing dwelling mounds and continuous dykes. From 1200 AD onwards, local and regional water authorities were founded in the Netherlands to control the large waterworks. The technical problems were solved by new technology: the ‘polder’ system of total endykement was introduced, at first based on gravity-flow drainage and later, when this technique was no longer adequate, on wind-powered pumping mills. By 1460, the windmill had become a common feature of polder drainage in the western peatland area of the Netherlands. Many shallow lakes, partly resulting from dredging peat, were pumped dry by windmills and converted into fertile croplands. The new technologies were spread by international experts from the sixteenth century onwards. Italian engineers recovered Combe Marsh near Greenwich UK, Dutch experts and workmen drained Plumstead and Erith marshes by the Thames. Humphrey Bradley from the southern Netherlands became Master

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of the Dykes of the French King Henry IV and supervised the draining of the Poitevin marshes north of La Rochelle. In 1626 the Dutchman, Cornelius Vermuyden, started the drainage of Hatfield Chase and, from 1631 onwards, he worked on the Great Level (Purseglove 1988). Dutch expertise was also employed in the drainage of major peatland areas in Germany in the early 1600s (Drews et al. 2000), including the Havelländische Luch and the Oderbruch (Witzel et al. 1999). Inspired by what he had learned during his stay in the Netherlands, Czar Peter the Great (1689–1725) introduced organised drainage of wetlands for agriculture in Russia (Paavilainen and Päivänen 1995). Despite the dense tree cover and the difficulty in land clearing, the fertile alluvial floodplains were the first in the southern United States to be converted for agricultural crops (Kellison et al. 1998). Intensive agricultural exploitation of wetlands began when South Carolina planters imported a suitable variety of rice from Madagascar in the 1690s. Even more significant, however, was the importation of slave labour. Africans brought the necessary work force and experience in rice cultivation. They also had stronger resistance to diseases such as malaria and yellow fever that had gained footholds in the flooded fields of the American South by the mid-1600s (Vileisis 1997). Following attempts to control water flow, first by private enterprise and then by public agencies, especially the US Army Corps of Engineers, the forests in the major alluvial floodplains were increasingly cleared for agricultural crops (Clark and Benforado 1981). By the 1930s, only half of the original forests remained. During the following 50 years, conversion continued at an accelerated pace, reducing the area from 4.8 to 1.7 million ha. Conversion was especially rapid during the 1960s and 1970s when the price for farm crops, especially soybeans, reached unprecedented levels (Turner et al. 1981; Kellison et al. 1998). One of the largest continuous mire areas of the former Soviet Union, the mires of Polesia in Belarus and Ukraine, were largely drained in the 1970s and 1980s (Bambalov 1996; Belokurov et al. 1998).

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Often the benefits of peatland drainage for agriculture may be only temporary. In drier climates especially, peat soils degrade rapidly and irreversibly after intensive drainage (Brandyk and Skapski 1988; Stegmann and Zeitz 2001) resulting in a drop in agricultural productivity (Succow 2001). Other negative side-effects include increased emissions of greenhouse gases and nutrients, decreased capacities for water storage and purification, and a loss of biodiversity (Bambalov et al. 1999; Schultz-Sternberg et al. 2000; Trepel 2000; Succow and Joosten 2001). Subsidence and oxidation after drainage induce a continuous lowering of the surface, making gravity drainage eventually impossible. The rate of subsidence and oxidation increases with increasing drainage depths and higher temperatures (Andriesse 1988; Eggelsmann 1990; Maltby and Immirzi 1993). In large areas of central Europe, this process has advanced to such an extent that the costs of pumping and dyke maintenance can no longer be met by agricultural revenues. The state of Mecklenburg-Vorpommern (FRG) decided in 2000 to remove agriculture from 75 000 ha of peatlands that can no longer be drained economically (Berg et al. 2000). Part of this area already has flooded ‘spontaneously’ (Succow and Joosten 2001). Peatland erosion by grazing Mires on slopes are very susceptible to erosion when their plant cover is damaged and the peat mass is exposed to frost, drought, rain and wind. Blanket mires in precipitation-rich areas are sensitive systems that can easily be damaged by grazing. Grazing is often accompanied by regular burning of the vegetation, which reduces the protective plant cover, increases the volume of surface runoff water, and increases the time during which the bare peat is exposed to frost and wind action. In former times, the higher blanket mires in the British Isles were used only in spring and summer for low-intensity grazing of cattle, sheep and goats. Over-grazing became a problem after sheep replaced cattle as the major herbivore and

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Human Impacts seasonal grazing was superseded by year-round grazing. The time of this changeover varied from around AD 1550 in the southern Pennines to around AD 1900 in Shetland (Tallis 1998). In every blanket mire area of Britain, sheep numbers further increased in the twentieth century, particularly since 1980 when the European Union introduced a sheep meat subsidy, the ‘ewe premium’ (Douglas 1998; Fuller and Gough 1999). At present, severely eroded peat, largely a result of over-grazing (Tallis et al. 1997), is found in all blanket mire regions of the British Isles. Only 4000 km2 of the blanket mire resource is still in a near-natural state, while 3500 km2 is eroded and a further 3500 km2 has been afforested (Tallis 1998). Mountainside vegetation has been severely denuded by sheep and, in places, the peat cover has totally disappeared. Severe losses of peat during the rainy winter months leads to peat siltation, clogging of spawning gravels, decreasing water clarity and reduced fish survival rates (The Heritage Council 1999). In the mountainous continental areas of Central Asia, grazing by livestock quickly damages the sloping peatlands that depend on surface-flowing water. Since the break-up of the Soviet Union, the peatlands in the IssykKul-area of Kyrgizstan, formerly used as meadows and hayfields by collective farms, were divided among the villages. Nowadays ‘privately owned’ dairy cows are grazing in the peatlands in high numbers and concentrated herds. During winter, they are joined by other cattle and horses that in summer graze higher mountain areas. As long as the peatlands are frozen (November–March), no harm is done. Intensive trampling from spring to autumn, however, has changed the vegetation of Carex compacta, C. orbicularis and C. pamirensis, to a community of Carex divisa hummocks and hollows with Carex orbicularis. The hummocks degrade by oxidation, whilst the areas between the hummocks are increasingly damaged by trampling. Peat particles flush away and accumulate as peat mud in the lower parts of the peatlands. Most of the peat was lost within a few years, leaving a mineral soil on which only a few remnants of the former peat layer may still

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be visible. Because of the diminishing forage, the animals take refuge in neighbouring, less damaged, peatland areas that in turn are being irreversibly destroyed (Heinicke 1999). Lesotho is currently one of the most severely eroded countries in the world (Calles and Kulander 1996). The montane ecosystems were relatively unaffected by humans until the arrival, around 200 years ago, of the Basuto, who brought permanent agriculture and livestock rearing. They settled mostly in the lowlands, but practised transhumance, that is they took their herds into the mountains for the summer grass and returned to the lowlands for the winter (Meakins and Duckett 1993). Around 1890 population pressure and shortage of pasture forced the Basuto into the high mountains, and now even the highest peaks are grazed in summer, with the additional use of mountain valleys in winter. Over-grazing of the mountain pastures forces the stock onto the boggy and marshy areas. Loss of vegetation cover on many of the warmer slopes in the subalpine and lower alpine belts (i.e. 1400–3000 m) has led to increased runoff and erosion, with constant silt deposition on the wetlands (Morris and Grab 1997). Erosion gullies and rills concentrate water flow on to the wetlands and deep, narrow drainage channels have formed in many of them. Trampling by cattle and massive-scale burrowing by Sloggett’s rat (Otomys slogetti) and Sclater’s golden mole (Chlorotalpa sclateri) in the drying peat accelerates the process of peat oxidation and erosion (Meakins and Duckett 1993). Remote sensing and field surveys indicate that approximately 65% of the mires in Lesotho are damaged in one way or another (Schwabe 1995).

The role of fire in wetland reclamation and management General Natural fires in wetlands have been reported worldwide (Brown 1990; Kangas 1990; Paijmans 1990; Kuhry 1994; Frost 1995; Zoltai et al. 1998). Fires affect wetlands in different ways, depending on their frequency and severity, the amount of

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organic matter or peat, and soil moisture conditions. Fires may change the rate of litter mineralisation and actually add structural diversity and biodiversity (Brown 1990), for example when lakes and pools originate from burned out holes in the peat. Some wetland species are adapted to, and even depend on, regular fires, such as the North American black spruce (Picea mariana) and pond pine (Pinus serotina), whose serotinal cones release their seeds after fire (Kangas 1990). The use of fire in agriculture is a widespread practice worldwide (cf. Steensberg 1993; Larsson 1995; Goldammer et al. 1997), including in wetlands. In swidden (slash and burn) agriculture, it is an effective way of clearing wild areas for crops. To maintain hunting or grazing ground, and in reed bed management, fire is used to clear old growth and to enhance new growth of young plant materials that are more palatable to wildlife and cattle, or more suitable for industrial purposes (Wirawan 1993; Hawke and Jose 1996). Fire clearance and the benefit of ash as a fertiliser have been known in Chinese tradition

Box 30.1

Law against moor-burning from King James I enacted in 1610 (After Evelyn 1661)

Anno vii. Jacobi Regis. An Act against burning of Ling, and Heath, and other Moor-burning in the Counties of Yorke, Durham, Northumberland, Cumberland, Westmerland, Lancaster, Darbie, Nottingham and Leicester, at unseasonable times of the year. Whereas, many Inconveniencies are observed to happen in divers Counties of this Realm, by Moore-burnings, and by raising of fires in Moorish grounds and Mountaneous Countries, for burning of Ling, Heath, Hather, Furres, Gorsse, Turffe, Fearn, Whinnes, Broom, and the like, in the Spring time and Summer-Times: for as much as thereby happeneth yearly a great destruction of the Brood of Wild-fowle, and Moor-game, and by the multitude of grosse vapours, and Clouds arising from those great fires, the Aer is so distemper’d, and such unseasonable and unnatural storms are ingendred,

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at least since the twenty-first century BC (Steensberg 1993). In China, Shi Chi (sixth century AD) used the phrase ‘Tilling with fire and weeding with water’, referring to wet-rice cultivation, where the rice was planted in fields full of water that impeded the growth of weeds. When the rice had been harvested, the stubble was burned to enrich the soil (Steensberg 1993). A charter from 1246 from Devonshire, UK, gives a clear reference to peatland burning (Dodgshon and Jewell 1970). In the Finnish peatlands of Savolax, Oesterbotten and Karelia, peatland burning for rye cultivation was practised extensively around 1650 (Schreiber 1912; Steensberg 1993). The burning of heath and moors to improve grazing goes back at least to the sixteenth century under the name of ‘muirburn’ (Fenton 1970). That this practice led to severe environmental problems is illustrated by a law against moorburning ‘at unseasonable times of the year’ that King James I (1603–1625) enacted in 1610 (Evelyn 1661; see Box 30.1). More recently burning is an essential part of moorland management for red

as that the Corn, and the Fruites of the Earth are thereby in divers places blasted, and greatly hindered in their due course of ripening and reaping. Be it therefore Enacted by our Sovereign Lord the Kings most excellent Majesty, …. That from, and after the last day of July next ensuing the end of this present Session of Parliament, it shall not be lawful for any Person or Persons whatsoever in the Months of April, May, June, July, August, and September, nor in any of them, to raise, kindle, or begin; or to cause or practise to be raised, kindled, or begun any fires or Moor-burnings in the said Counties of York, Durham, Northumberland, Cumberland, Westmorland, Lancaster, Darby, Nottingham, and Leicester, or in any of them, for burning of Ling, Heath, Hather, Furs, Gorsse, Turffes, Fearne, Whinnes, Broome or the like.

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Human Impacts grouse (Lagopus lagopus scoticus; Hobbs and Gimingham 1987). That this practice eventually could lead to a complete loss of the wetland resource was already expressed in the 1790s in an account from Lanarkshire (Scotland): ‘Moss, of this kind, repeatedly burnt, becomes thinner and thinner, till it disappears almost or altogether, and leaves the clay, that was once three or four feet down, on the surface. Some hundreds of acres have been converted in this manner from moss to make arable land’ (Fenton 1970). Peatland fires may lead to the ignition of the peat layers and a consequent extreme degradation of the ecosystem (cf. Maltby 1980). Peatland fires are difficult to extinguish and may last for many months. The depth and areal extent of such fires depend on the oxygen availability, the moisture content and the presence of cracks in the peat (Ellery et al. 1989; Maltby et al. 1990; Grundling et al. 1998). Two case studies will illustrate the parallel backgrounds of massive environmental damage due to peatland fires, one from the nineteenth century that led to the opening up of the last extensive peatlands of north-western Europe, the other from the late twentieth century in South-east Asia.

Buckwheat fire cultivation The technique of buckwheat fire cultivation was brought from the Netherlands to north-western Germany around 1700 (Schreiber 1912; Ahlrichs 1987). In the Reclamation Edict of King Frederic the Great (1765), it was introduced as the basis of colonisation of the East Frisian bogs (Schneider and Stoller 1926). The culture expanded rapidly in Germany during the eighteenth and nineteenth centuries as large expanses of mires, considered ‘idle areas’, were available (Westerhoff 1936). Numerous people, mostly without any means of subsistence, ‘often tramps and other rabble’ (Tacke and Lehmann 1912), were settled as colonists in the wild mires (Lehmann 1940). Buckwheat fire cultivation was the only means of exploiting the extensive uninhabited peatlands without transport ways, fertilisers and financial resources (Schreiber 1912). To prepare the mire for buckwheat cultivation, the bog was superficially drained, the surface was cut by hand or ploughed in autumn, and the sods – dried over winter – were kindled the next May, if the weather allowed (Freese 1789; Figure 30.1). The burning destroyed the unwanted vegetation, its seed bank and root system, it improved the structure of the peat soil and somewhat

Fig. 30.1 Moor burning in Germany. (Collection Hans Joosten.)

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neutralised its acidity, released nutrients from peat and biomass or changed them into forms better available for the subsequent crop (Schreiber 1912; Von Seelhorst 1914). In the still-warm mixture of ashes, charcoal and peat, the buckwheat (Fagopyrum esculentum) was sown (Venema 1855; Nöggerath 1875; Göttlich 1990). More sporadically, rape, potatoes, rye and oats were cultivated (Van Hall 1847). Around 1870, buckwheat fire cultivation may have covered an area in north-western Germany and the adjoining Netherlands of around 100 000 ha annually (Schreiber 1912; Lehmann 1940; Foorthuis 1991; Bieleman 1992). The crop was initially received with enthusiasm. ‘No cereal pays the labour richer than buckwheat’, wrote Freese (1789). Coolen (1874) described the extensive fields of flowering buckwheat: ‘as a bright white sheet, that lays there spread out and from where vapours and odours ascend, that caress and refresh the senses like a refreshing balm. The eye rejoices in this clear wedding-dress, the ear in the buzzing of the bee army that strikes up a jubilation, glad as it seems to find Man here in this flower festival as a rival in the field of industry’. Buckwheat could indeed supply rich yields on newly burned peatlands. But it was also a ‘lottery crop’ (Tacke and Lehmann 1912; Lehmann 1940): the harvest was regularly lost because of the frequent night frosts in the peatlands in summer or because of too much rain (Grisebach 1846). Its largest drawback, however, was that the yields decreased every year. After 6–7 years the soil was so depleted in nutrients that other fields had to be made (Van Bemmelen 1871; Gröninger 1910). Consequently, only some 15–20% of the total peatland area could be under simultaneous cultivation (Grisebach 1846; Van Bemmelen 1871; Bersch 1909). On the old fields, a fallow period of at least 25–30 years was necessary (Von Seelhorst 1914), but even then yields diminished continuously (Schneider and Stoller 1926). Other disadvantages gradually became clear. The drainage ditches had to be deepened continuously

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because of peat subsidence and oxidation. Deep fires could ‘burn the peatland to death’: the surface changed in to a dusty, pulverised mass that could not be wetted, and was easily eroded by the wind (Bersch 1909). The extensive drainage of the raised bogs caused inundations in the surrounding areas of mineral soils (Westerhoff 1936). Many negative environmental effects were observed. Nöggerath (1875) pointed out that an annually burned area of 50 000 acres caused atmospheric emissions ‘exceeding 1800 million pounds’. The ‘Moorrauch’ (moor smoke) spread over large areas (‘the whole of Germany notices when our peatlands burn’, Kutzen 1880) and even reached Hungary and the South of France (Van Bemmelen 1871; Tacke and Lehmann 1912; Figure 30.2). The moor burning was thought to generate wind, it destroyed bird nests and animals and led to many accidental fires (Stemfoort 1847). The smoke haze would drive away rain and thunderstorms and ‘disturb the free passage of the sunlight’. The haze was cold, led to night frosts, and was thought to have a negative impact on the health of humans, animals, vegetation and crops (Grisebach 1846; Nöggerath 1875; Prestel 1903; Schreiber 1912; cf. Davies and Unam 1999). A massive resentment grew. A hundred years after its official introduction in Germany, buckwheat fire cultivation was seen as ‘a hardly bearable barbarian culture’ (Van Bemmelen 1871), as predatory cultivation of the worst kind (Bersch 1909) and as a complete failure. At the end of the 1860s, villagers near the peatland settlements started to protest against the buckwheat fire cultivation (Van Bemmelen 1871). An overall ‘Verein gegen das Moorbrennen’ (Society Against Peatland Burning) was founded in 1870 by citizens of the frequently haze-burdened city of Bremen (Buchenau 1889). Results quickly followed: the Hanoverian government refused to issue new leases for ‘Brandland’ in the Gifhorner Moor region, an example followed by the regionally active Society of Braunschweigian Capitalists, and the grand duchy of Oldenburg generally forbade moor and heathland burning between the 1st of June and the 1st of September. The positive effects of these measures for the Rheinland area,

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Fig. 30.2 Map of the distribution of moor haze stemming from buckwheat fire cultivation in north-western Germany in the years 1848, 1857 and 1863. (After Prestel 1903.)

some hundreds of kilometres to the south, were reported immediately (Nöggerath 1875). A complete prohibition was, however, not yet possible because a large population depended solely on fire cultivation. Therefore the Society Against Peatland Burning worked towards development of alternative methods of bog cultivation. Moved by the discussions, the Prussian Minister of Agriculture called a Zentral-Moor-Kommission into existence (1876) with a primary aim to ‘research to make soil utilisation in the peatland districts more productive’ (Baden 1976). This resulted in the foundation of an experimental

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station, the ‘Preußische Moor-Versuchsstation’ in Bremen (1877), and the ‘Society for the Stimulation of Peatland Cultivation in the German Empire’ in 1883 (Schreiber 1912; Baden 1976; Cordes 1977; Kuntze 1977). Eventually a less destructive, and agriculturally more reliable, bog exploitation technique was found in the ‘German bog cultivation’ (Deutsche Hochmoorkultur), a practice dating from around 1750 but further developed and stimulated by the Moorversuchsstation (Brüne 1931). The culture included systematic drainage and 40 cm deep ploughing of the virgin peatlands, followed by

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the application of chalk and fertiliser. Buckwheat fire cultivation was still used as a preparation for further reclamation until it was forbidden by the peatland conservation law of 1923 (Hasekamp and Wieking 1976). The areas cultivated by the Deutsche Hochmoorkultur were excellent as grassland, but as arable land they lost 2–3 cm a−1 peat by oxidation, resulting in a major loss of productivity when the strongly humified Sphagnum peat was exposed (Hasekamp and Wieking 1976). After the Second World War, many shallow (up to 2 m deep) peatland areas were reclaimed by deep-ploughing to provide land for part of the displaced population driven westward from beyond the Oder-Neisse line (Schoenberg 1970). The deeper areas have since remained in grassland cultivation, but are becoming increasingly abandoned because of soil degradation and drainage problems (Blankenburg 1999), indicating that agriculture on drained peat soils is not a sustainable activity. South-east Asian peatland fires A major concentration of the global tropical peatlands can be found in South-east Asia. Indonesia possesses the largest area of peat in the tropics (up to 27 Mha), although the estimates for the extent of this resource vary considerably (Maltby and Immirzi 1993; Maltby et al. 1996; Rieley et al. 1996; Rieley and Page 1997). The development of the Indonesian peatland swamps has been linked with plans to redistribute people from the over-populated islands of Java, Madura, Bali and Lombok (with around 700 people km−2) to the outer islands where there are only some 30 people km−2 (Manshard and Morgan 1988; Scholz 1988). State organised transmigration programmes (‘transmigrasi’ in Indonesian) were first attempted by the Dutch in the then Netherlands East Indies in 1905, both to alleviate the population problem in Java and Bali and to provide a labour supply for the estates in Sumatra (Bahrin 1988). Between 1905 and the end of the colonial period, resettlement brought a yearly average of 6300 people

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from Java into the Sumatran province of Lampung (Zimmermann 1980). After independence (1947), the Indonesian governments attempted to give new impetus to transmigrasi. The greater sums of money available from petroleum earnings enabled the total number of transmigrants to rise to 835 000 in the period 1950–1973 (Manshard and Morgan 1988). Apart from the governmental transmigrasi programmes, an even larger flow of spontaneous pioneer settlers has emerged, outnumbering the state directed migrants by several times (Scholz 1988). The acid peatland swamps were generally considered ill-suited to agricultural use (Driessen and Soepraptohardjo 1974; Manshard and Morgan 1988), but the absence of competing land use claims, and the suitability of these for a combination of tidally-irrigated wet-rice cultivation and coconut planting, made the coastal wetlands into a focal area for transmigration (Scholz 1988). Wet-rice cultivation is generally restricted to a strip close to the river. At the points where mineral soils gradually give way to peat soils, there are land use changes in favour of coconut cultivation. Behind the coconut strip the swamps usually become too deep for any kind of land use and are left uncultivated. Spontaneous swamp colonisation had already begun in around 1900 in Sumatra (Scholz 1988; Figure 30.3), to be followed by larger scale systems in South and Central Kalimantan in the 1950s and 1960s (Brookfield et al. 1995). In the 1960s, when various sawah (irrigated rice field) intensification and expansion programmes on Java had failed because of the limited availability of land with adequate irrigation facilities, Indonesia was confronted by a serious rice shortage, making the country the world’s largest importer of rice. The tidal swamps were looked to for a ready solution, and a national programme started in 1968. With tremendous inputs of capital and machinery, large sections of the lowland swamps were turned into farm land for wet-rice cultivation (Scholz 1988). By the late 1970s more than 800 000 ha of coastal wetlands in Indonesia had been opened up for rice production, principally in Kalimantan and Sumatra.

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Fig. 30.3 Swamp reclamation for settlement projects in the Batanghari delta, Sumatra, Indonesia. (Changed after Scholz 1988.)

The introduction of high-yielding rice varieties and major programmes of fertiliser subsidy, irrigation improvement and rural credit led to the attainment of Indonesian self-sufficiency in rice in the mid-1980s (Chang 1993; Brookfield et al. 1995). At the start of the intensification programme, rice production averaged around 11 million tonnes. By the mid-1980s, it had more than doubled to over 25 million tonnes, and by 1989 it had reached 30 million tonnes (Fox 1993). This increase in rice yield in the existing sawahs, however, levelled off in the following years. Furthermore the best sawah lands in Java were gradually lost to rapid industrialisation and urbanisation (Notohadiprawiro 1998). After 1993 the country again had to import rice in increasing

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quantities (to over 3 million tonnes in 1998). In order to restore Indonesia to self-sufficiency in rice production, President Suharto decided to compensate for the lost rice land by claiming an equivalent area in Central Kalimantan (Rieley 1999b). By presidential decree of 1995 a ‘mega-project’ to develop one million hectares of peatland for food crop production in Central Kalimantan was initiated, aimed at the production of two million tons of rice per year. It was thought that the high investment costs for irrigation infrastructure could largely be saved by choosing peatlands where the necessary water was already available. The total area of the mega-project was 1 457 100 ha of which peat-swamps occupied

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919 060 ha, comprising 41% of the total peatswamps in Central Kalimantan (Notohadiprawiro 1998). The project involved the wholesale destruction of forest in order to clear the area prior to conversion to plantation agriculture for rice cultivation (Rieley 1999a). The Parent Primary Canal (110 km long, 25 m surface width, 15 m bottom width, 6 m deep) was dug, branching into 1129 km of main primary canals that in their turn branch into 964 km of secondary channels, 900 km of tertiary and 1515 km of quaternary channels (Figure 30.4). The canals were mapped out and excavated without factual knowledge of the hydrotopography and without taking the hydraulic and hydromorphological properties of peatlands into account. The Parent Primary Canal, for example, was intended to supply fresh riverwater to flush the peatlands, but as the contours of the peatland surface between the rivers are domed, the opposite occurs, and the peat drains towards the river. Many stretches of canals are draining too deep, resulting in subsidence, desiccation, peat oxidation, acidification by sulphide oxidation and fire danger (Notohadiprawiro 1998). The biomass residues from these and similar land-clearings in recently drained peat-swamps in Sumatra and Kalimantan provided the fuel for immense, uncontrollable fires in 1997 and, to a smaller extent, in consecutive years. Liew et al. (1999) mapped areas affected by the forest fires in South Kalimantan and found a total of 552 000 ha of land had been burned in 1997 out of the 3.6 Mha area surveyed. The fires were lit by companies and farmers as a cheap way of clearing land for agriculture and commercial plantations of palm oil, rubber and timber. An El-Niño drought exacerbated the situation. Much of the surface-drained peat that was supposed to support crops burned away (Murdiyarso and Lebel 1998). Large areas of South-east Asia were blanketed with smoke leading to immense environmental problems (cf. Wirawan 1993; Murdiyarso and Lebel 1998). There were 1000 smog-related deaths, and in 1997 more than 40 000 people suffered respiratory ailments. In neighbouring Malaysia, Singapore

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and Brunei, schools had to be shut down, airline flights cancelled and people wore face masks to protect their lungs from the polluted air. Apart from the immediate impacts of fires and haze on people’s livelihoods and on biodiversity in Indonesia, the transboundary haze had widespread effects on health, transportation and tourism industries in the neighbouring countries, especially in Malaysia, Singapore and Thailand. Levine (1998) estimated the emissions in 1997– 1998 to have been 85–316 Mt CO2, 7–52 Mt CO, 0.2–1.5 Mt NOx, and 4–16 Mt particulate matter. Page et al. (2002) estimated the release to the atmosphere in 1997 as a result of burning peat and vegetation in Indonesia as 810–2570 Mt C. This is equivalent to 13–40% of the mean annual global carbon emissions from fossil fuels, and contributed greatly to the largest annual increase in atmospheric CO2 concentration detected since records began in 1957. The haze and fires during August–December 1997 cost the region an estimated US$ 4.4 billion (Glover and Jessup 1998). Just two and a half years after its commencement in January 1996, it became apparent that the project would never achieve its objectives and the President of Indonesia, H.J. Habibie, reversing the decision of his predecessor Suharto, stopped the project. By this time around three trillion Rupiah (500 million US dollars) had been invested, but only some 3% of the plan had entered the final stage. Some 50 000 ha had been fully equipped with water controlling devices and were settled by 13 500 families of local people and transmigrants (Notohadiprawiro 1998). Most of the forest within this area, however, had already been destroyed or degraded irreparably (Rieley 1999a). An official embargo on further logging within the area of the Mega Project was imposed by the Ministry of Forestry on 1 July 1998, and the logging licenses of 29 companies who had failed to report their suspected roles in starting the fires were revoked (Rieley 1999a). Environmentalists claimed this represented only political lip service, since logging firms in Indonesia often operate without permits anyway (Kriner 1999). The channels provide easy entry in the formerly inaccessible peat-swamp forests and enable

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the illegal loggers to transport wood by floating logs to the nearest river. As a result, all remaining timber is being removed and, in the process, debris is set alight and the surface peat catches fire. Illegal timber extraction has increased to such a scale that the forest may never recover (Rieley 1999a). On 13 July 1999 President Habibie issued Decree 80/1999 in which the failure of the Mega Rice Project was recognised. The one million hectare area has been rolled up into a new 2.8 million hectare economic development zone with the focus on land conversion to food crops and plantations, especially oil palm and rubber (Rieley 1999b). Although the land should be developed and managed according to wise use and sustainability principles, these concepts remain unclear and ambiguous, as the main emphasis is on economic development. The decree reveals a lack of understanding of the functioning and values of tropical peatlands and peat-swamp forests. Areas designated for hydrology, peat and wildlife protection are fragmented and are not sustainable because of their proximity to large developmental areas. Wildlife conservation in large areas of wilderness is not mentioned at all (Rieley 1999b).

F OR E S T R Y Forestry has led to wetland destruction, both by cutting and by planting of trees. Deforestation of wetlands for agriculture has been widely practised all over the world. In Europe this has resulted in the loss of virtually all floodplain forests. Thirtytwo percent of the formerly forested peatland area of peninsular Malaysia is currently utilised for agriculture (oil palm, rubber, coconut) and mixed horticulture (Ambak and Chye 1996). In the United States, forested wetlands have been lost at rates five times higher than nonwetland forests (Brinson 1990). During the 1700s, cypress (Taxodium distichum) was the principal cash product for most colonists of the lower Mississippi Valley. Large-scale commercial logging of cypress began when the Homestead Act of 1866, which declared swamp lands unfit for

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cultivation and unavailable to private individuals, was repealed by the Timber Act of 1876. By the close of the nineteenth century over 7 million m3 of bald cypress had been logged in Louisiana (Conner and Buford 1998). In the Carolinas, the introduction of steam dredging and railroad logging in the 1850s led to the harvest of virtually every known stand of Atlantic white-cedar (Chamaecyparis thyoides) (Conner and Buford 1998). Some mires support tree growth naturally, but waterlogging usually prevents an economic level of wood production. However, black spruce (Picea mariana) is harvested in significant volumes from forested mires in Canada, especially Ontario and Quebec, as a source of fibre for the pulp and paper industry (Paavilainen and Päivänen 1995). Similar tree harvesting on virgin forested mires is hardly practised in Europe. The idea of draining peatlands to promote tree growth was first published in Sweden in 1773 (Paavilainen and Päivänen 1995). Peatland drainage in Russia began in 1775 with the drainage of a 970 ha area of Picea and Alnus forest along the Ochta River near Moscow, which was partly used for forestry and partly for grassland (Konstantinov et al. 1999). Since then more than 5 million ha of forested wetlands and mires have been drained in Russia. Systematic drainage for forestry started in the Baltic and Fennoscandian states in the nineteenth century (Paavilainen and Päivänen 1995). Most peatland drainage for forestry has taken place in Finland and Russia since about 1950, with Russia reaching an annual maximum of 260 000 ha in 1960, and Finland of 295 000 ha in 1969 (Paavilainen and Päivänen 1995; Päivänen and Paavilainen 1996). Finland and Russia are responsible for 60% of the global mire losses to forestry (Table 30.3). On over 1 million ha of the drained area in Russia, the drainage canals no longer function because of neglect, activities of beaver or drainage-obstructing infrastructure (e.g. roads and pipelines) and the land reverts to peatland (Konstantinov et al. 1999). Vomperskij (1999) states that half of the approximately 6 million ha of drained peatland are nowadays back in their ‘original situation’.

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Human Impacts Table 30.3 Peatlands (and partly forested wetlands) drained for forestry. (After Paavilainen and Päivainen 1995.) Km2 Finland Russia Sweden Norway Estonia Latvia Lithuania Belarus Poland Germany United Kingdom Ireland P.R. of China USA Canada Total

59 000 38 000 14 100 4 200 4 600 5 000 5 900 2 800 1 200 1 100 6 000 2 100 700 4 000 250 148 950

Drainage of new areas has practically stopped (Konstantinov et al. 1999). In the boreal zone of North America, much less peatland has been drained. Current demands from the paper industry for a continuous supply of fresh black spruce (Picea mariana) may change this in the near future. In the southeastern USA, the pocosins (‘swamp-on-a-hill’ in the Algonquin language) have suffered heavily from forestry. Pocosins are nutrient-poor, forested or shrub peatlands and comprise approximately 50% of North Carolina’s freshwater wetlands. Of the original 1 million ha, only some 30% have remained in their natural state (Hartmann et al. 1994). Federal economic incentives (cost sharing and tax concessions) have encouraged the conversion of these freshwater wetlands to forestry. Drainage of treeless mires and afforestation with Scots pine (Pinus sylvestris) have been quite intensive in Europe since the 1950s, for example in Finland (a total of 235 000 ha in 1979) and Norway. In Atlantic Europe, afforestation has concentrated on the blanket mires of Scotland, and the north of England, Wales and Ireland (Anderson 1997; Tallis 1998). Deep

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furrows were ploughed to remove the water from the uppermost peat layers. Often exotic species such as Picea sitchensis and Pinus contorta were planted (Paavilainen and Päivänen 1995). In Britain blanket mire afforestation was stimulated by a combination of state grants and tax incentives – as is shown by the increasing importance of the private sector in afforestation in Britain during the 1980s – and, since then, has rapidly decreased due to the change in tax laws, consequent of strong protests (Lindsay et al. 1988; Keatinge 1990). Apart from changes in species and ecosystem diversity (cf. Ratcliffe and Oswald 1987; Vasander 1990; Eurola et al. 1991), the environmental effects of peatland forestry, especially consequent to digging and clearing ditches, may include: • increase in peak flows, especially after clear cutting, which may lead to rapidly varying salinity in estuaries and lower production of shrimp, finfish and oysters; • increased runoff of suspended solids leading to the silting up of watercourses; • increased runoff of nutrients (phosphorus, potassium and organic nitrogen), especially after fertilisation, leading to eutrophication and increased algal production in nearby streams and estuaries; • changes in pH in watercourses downstream, generally leading to acidification; • decrease of economically important species, for example in the boreal zone wild cranberry (Oxycoccus palustris) and (after an initial positive effect) cloudberry (Rubus chamaemorus) (Hartmann et al. 1994; Paavilainen and Päivänen 1995; Sharitz and Gresham 1998). With respect to the peatland’s carbon balance and its radiative forcing, the effects of drainage for forestry are complicated and strongly dependent on the time frame chosen (cf. Crill et al. 2000; Joosten 2000).

PEAT EX T R ACT ION Peat extraction originated in early times as shown by archaeological finds of peat sods, peat

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balls, peat ashes,and peat pits dating from the Neolithic in northern Germany (Dieck 1983), the Bronze Age for example in the northern Alps, the Pre-Alps, and northern Germany (Dieck 1983), the pre-Roman Iron Age in Denmark (Rasmussen 1970) and Scotland (Carter 1998; Dickson 1998), and from Roman times in the Netherlands, Germany and England (Schütrumpf 1951, 1958; Jankuhn 1958; Clason 1963; Hall et al. 1980). Peat spades date from the Iron Age (Manning 1970; Averdieck and Schneider 1977). Many objects, including human remains, have been deliberately buried in peat, and were afterwards overgrown with newly formed peat (Glob 1965; Coles and Coles 1989; Van der Sanden 1990). Peat has been extracted as a raw material for a large range of domestic, industrial and environmental products, varying from nappies (diapers) to raw material for the chemical industry, for medicine (balneology), as litter material in cowsheds and chicken farms, for sludge handling, composting, cleaning of wastewaters, biological air purification, and oil-absorption, and even for peat textiles (Joosten and Clarke 2002). Peat for energy The heating value (in MJ kg−1) of strongly humified peat is approximately half that of coal (Asplund 1996). Peat, however, has many advantages: it burns with a long flame, it forms only 1–3% ash, little soot, and the fire needs little draught. Its greatest drawback is its volume, since it needs six times the storage capacity of coal (Gerding 1995). In some regions, peat cutting for fuel has taken place since the Neolithic (Dieck 1983).The first historic account of peat usage for fuel is found in the Historia Naturalis of Gaius Plinius Secundus (Pliny the Elder, 24–79 AD), who described the habits of the German tribe of the Chaukes: ‘captumque manibus lutum ventis magis quam sole siccantes terra cibos et rigentia septentrione viscera sua urunt’ (Plinius, Hist. Nat. 16, I, 4): ‘and the mud kneaded with their hands they dry more by winds than by the sun and with the

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earth they heat their food and their flesh stiff by the north wind’. In Ireland, the eighth century law tract Senchus Mór covers rules for cutting peat and carrying it to the house when dry (McKennan 2000). In the second part of the ninth century, peat fuel was reported from the Orkney Islands (Niedner 1924). An Arabic traveller from around 973 AD (Crompvoets 1981) wrote about Utrecht (Netherlands) in his itinerary, ‘there is in their country no wood for heating, only a loam, that replaces wood. And they go in summer, when the waters have spread, to their meadows where they cut the loam in the form of bricks with axes. Everybody cuts as much of it as he needs, and then spreads them out towards the sun to dry. Consequently it becomes very light. When it is brought into the fire, it ignites, and the fire takes hold of it, like it takes hold of the wood, and it makes a large fire with mighty glow like the fire in a glass oven. When the piece is burned, it does not leave charcoal, only ashes’ (Jacob 1890). The Netherlands was the first country where large-scale peat extraction was practised. In 1650, the peatlands around Groningen satisfied 40% of the national energy demand (Gerding 1995). The first scientific book on peatlands and peat extraction (Schoockius 1658) was published in Groningen by university professor Schoock, who obtained his knowledge from his students, the sons of major local peat extractors (Van Dijk et al. 1984). The amazing prosperity of the Netherlands during most of the seventeenth century was based on the energy provided by ample peat supplies, which could be easily transported by boat in the flat and wet country (De Zeeuw 1978; Gerding 1995; cf. Unger 1984). Peat extraction for fuel (Figure 30.5) reached a new boom in 1850–1925. Peat in the Netherlands in 1859 supplied 7.5 gigajoules of energy per head, with a similar amount from coal. Afterwards the importance of peat for fuel decreased sharply. In 1939, peat supplied only 3% of the national energy demand (Gerding 1995), and large-scale peat extraction in the Netherlands stopped completely in 1992 (Joosten 1994).

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Fig. 30.5 Peat extraction systems in the north-eastern Netherlands. (After Gerding 1995.)

The energy use of peat currently accounts for about 50% of the total global peat extraction volume. Peat is important as a fuel in countries where other fossil fuels do not exist or where long transport distances favour the exploitation of local peat resources (Asplund 1996). It is

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primarily used for centralised production of hot water and electricity and, to a minor extent, for private consumption (peat briquettes and some private fuel production, Asplund 1996). In the Central Asian steppe republics, like Kyrgizstan and Kazakhstan, local communities have taken

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Table 30.4 Peat extraction (10–6 tonnes) for energy (after Joosten and Clarke 2002). The figures are not strictly comparable, as different countries estimate weight in relation to different moisture content. In addition, the table combines milled and sod peat at different moisture content. Extraction volumes vary from year to year, depending on the weather (cf. unfavourable weather in 1998). Country

1990

1997

1998

1999

Russia Belarus Ukraine Estonia Finland Ireland Sweden Latvia Lithuania Total

6.0 3.4 1.3 0.0 5.8 7.5 0.0 0.3 0.8 25.1

2.9 2.7 0.6 0.2 10.1 4.0 1.4

1.9 2.0 0.6 0.2 1.5 4.3 0.2 0.1 0.2 11.0

3.7 3.1 0.5 0.6 7.5 4.7 1.1

0.1 22.0

0.4 21.6

up peat extraction for fuel again after the break up of the Soviet Union, because electricity and wood are not sufficiently available and coal is too expensive (Heinicke 1999). Finland, Ireland, Russia, Belarus and Sweden consume considerable volumes of peat for energy (Table 30.4). In Finland, peat extraction for energy consumes roughly six times more peat than is accumulating in the undisturbed mires of that country (Crill et al. 2000; Joosten 2000). In Ireland peat represents some 10% of the primary energy production (Fernandez Ruiz 1994), and new power plants with a combined capacity of 370 MW, that consume three million tonnes of peat each year, have been built to replace the older generation of peat-fuelled plants (O’Rourke 2000; Joosten and Clarke 2002). In several countries, a reorientation on peat was stimulated by the oil crisis in the 1970s. In 1974 the Finnish government accepted a programme to expand peat fuel supply to 20 million m3 a−1. In 1994, Finland attained a production record of 28 million m3, equivalent to 2.2 million tonnes of oil (Sopo and Aalto1996). The importance of peat as a fuel is expected to decrease in future, at least in Europe, because

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of the opening up of the electricity market and the environmental pressure to reduce greenhouse gases. Furthermore, in major peat extracting countries, peat will be virtually exhausted by the end of the first quarter of the new century (Hughes 2000). Salt making Next to fuel, salt making was for centuries one of the important peat-consuming branches of industry, since salt was an essential product for conserving food. In Roman times, salt extraction was common in the coastal area of north-western Europe, as is shown by archaeological evidence in western Belgium and adjacent areas of France and the Netherlands (Borger 1992; Gerding 1995). Salt could be obtained by evaporation of seawater (French ‘briquetage’) or by burning of salt-impregnated peat (Dutch ‘darink delven’) or saliferous plants. Since climatic conditions were not favourable for natural evaporation, and trees were scarce in the mires, most of the fuel had to be provided by turbaries (Overbeck 1975; Borger 1992; Gerding 1995). The Norfolk Broads originated as a result of this activity (Lambert et al. 1960). As the peat was often dug immediately behind the dykes or on the foreshore, this large scale peat utilisation weakened coastal defences considerably (Drews et al. 2000). Consequently, one of the oldest and largest poldered areas in the Netherlands, the Grote or Zuidhollandse Waard, flooded during the St Elisabeth flood in 1421 and was transformed into a new wetland that never was redyked: the present Biesbosch National Park (Lambert 1985; Borger 1992; Beenakker 1994). Agricultural and horticultural peat Agricultural and horticultural use is now the main purpose of peat mining across the world. Annual extraction at the end of the 1980s may have been 150–200 million tonnes, 95% of which was extracted by the former Soviet Union where large-scale agricultural production in collective and state farms used enormous quantities of peat

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Human Impacts Table 30.5 Peat extraction (10–6 m3) for horticultural and agricultural use (after Joosten and Clarke 2002). These figures may not be comparable as different moisture contents may be used. Country Belarus Russia Ukraine Czech Rep Estonia Finland Germany Hungary Ireland Latvia Lithuania Poland Sweden Norway Denmark UK USA Canada N Zealand Total

1990

1997

1998

1999

10.9 24.0 9.8 0.3 0.0 1.5 6.6 0.1 2.0 13.2 2.0 0.3 0.0

0.3 2.5 0.1 0.4 3.5 1.6 9.0

0.4 0.8 0.2 0.1 1.0 0.3 9.6 0.1 3.2 0.7 0.4 0.7 0.6

0.8 1.1 0.3 0.2 3.5 2.4 9.5 0.2 3.2 0.0 0.8 0.8 1.4 0.1

1.9 1.4 8.8 0.1 30.3

2.5 1.4 10.3

1.4 1.2 6.6 0.0 79.9

2.8

0.7 0.6 0.1 0.5 2.5 2.2 7.0 33.8

38.5

for soil improvement (Nutek 1992; Robertson 1994). Since 1989, peat extraction volumes in this part of the world have considerably decreased. Annual peat extraction in Lithuania was 2.3 million t in 1985 but just 250 000 t in 1995 (J. Karpavicius, personal communication). There has been a reversal of this trend in recent years due to increased exports to Western Europe (cf. Table 30.5). In the last 25 years, drastic changes have taken place in patterns of peat consumption in western and central Europe. In the beginning of the 1980s, peat was still used primarily for agricultural soil improvement and hobby gardening. This has rapidly changed, and professional horticulture is now the major consumer of peat. The value of Sphagnum peat in horticulture lies in a unique combination of properties which enable it to retain large amounts of water, entrap large volumes of air and hold large quantities of plant

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nutrients in readily available form. Furthermore, it has the advantage of low pH and nutrient content, which facilitate formulation of growing media for a wide range of applications by adding nutrients and other materials. Since 1980, increased use of horticultural peat has been observed in almost every country. Countries with major peat resources began to extract increasing volumes of peat, both for domestic horticulture and for export (Joosten 1995). As very few peatlands are left in the Netherlands and Germany, where domestic peat resources were once abundant, these countries now import peat in increasing quantities, for example from the Baltic states, and act as major redistributors of peat to all over the world (Joosten 1995). Because of these international connections and the diffuse use of peat in professional horticulture, anti-peat campaigns in the 1990s have barely influenced peat extraction volumes, although British and Irish environmental groups did succeed in reducing domestic peat use by private households (Joosten 1995; Shaw 2000). While peat has maintained its position as the leading material for growing media, alternative materials have emerged as substitutes but, at present, no equally risk-free alternative material is available in large enough quantities to replace peat in horticultural crop production. In the near future an increase may therefore be expected in the extraction and use of horticultural peat, resulting from higher demands and higher prosperity. Peat extraction is responsible for only 10% of the losses of the global mires (cf. Table 30.2), nevertheless peat extraction destroys ecosystems over large areas. Large scale peat digging requires the reduction of the water content of the peat from 95% to 90%, meaning that half of the total water storage in the peat body is removed. Drainage also leads to discharges of organic suspended solids varying from few tonnes up to about 30 tonnes km−2 a−1 (Sallantaus and Pätilä 1985). Migration of solid peat particles into watercourses can, however, largely be counteracted by ditch retainers, sedimentation in settling ponds and filtering (Selin 1996).

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CO N CL U S IO N S Outside the tropics, human exploitation has damaged 500 000 km2 of mires so severely that peat accumulation has stopped completely. In recent years, the rates of mire destruction are an order of magnitude greater than the mean annual mire expansion rate during the Holocene. The global peat resource shows a similar picture. Peat extraction for fuel and horticulture is presently responsible for oxidative peat losses of some 15 × 106 t a−1 C (excluding carbon emissions from peatlands drained for and by extraction, cf. Joosten and Clarke 2002), while agriculture and forestry release 100–200 × 106 t a−1 C (Immirzi et al. 1992; Vomperskij 1999). As global peat accumulation is about 40–70 × 106 t a−1 C (cf. Gorham 1991; Immirzi et al. 1992; Maltby and Immirzi 1993; Clymo et al. 1998; Turunen et al. 2002), the world’s peatlands have changed from a carbon sink to a carbon source. The global peat resources are currently decreasing at about 0.05% per year (cf. Armentano and Menges 1986; Maltby and Immirzi 1993; Botch et al. 1995; Lappalainen 1996; Vompersky et al. 1996; Joosten and Clarke 2002). The challenge for the twenty-first century is to develop exploitation techniques that enable a sustainable use of peatlands, while preventing negative environmental side-effects. Peatland exploitation must therefore be redirected from its present focus on the consumption of fossil resources to the utilisation of the wider range of benefits which can be provided by intact wetland ecosystems. An important role in reducing the threat for virgin mires must be played by the restoration of the many degraded and often abandoned peatlands. All translations in the text are by Hans Joosten.

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Section VI Wetland Restoration: Making Wetlands Work Again

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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31

Introduction – Re-Establishment of Wetland Functioning E DWARD MALTBY

Institute for Sustainable Water, Integrated Management and Ecosystem Research, University of Liverpool, Liverpool, UK

IN T R O D U CT ION Wetland restoration is not just aquatic gardening. The number and scale of wetland restoration projects is growing exponentially worldwide. Van der Valk (Chapter 32) concludes that established ecological theory has been used only sparingly in the design and implementation of past restoration projects, however, this situation is changing rapidly as a whole new discipline is developing in response to the widespread degradation of wetland ecosystems and increased awareness of the benefits from their restoration. The term ‘restoration’ itself can cover a range of emphases and associated activities. Repair, rebuilding, replacement, enhancement, rehabilitation, creation and mitigation are just some examples of the terminology often used. In Chapter 32, Van der Valk suggests there are just two broad categories of project: restoration and creation. A further distinction is made between ‘historic’ projects, which aim to replicate previously existing wetland types, from ‘functional’ projects, which aim to establish the properties and conditions for processes to take place in support of particular wetland functions. In the first case, success might be measured by the plant and animal species established, while in the second, it might be the amount of floodwater stored or

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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nutrients retained, thus contributing to flood peak reduction and water quality respectively. The fundamental distinction between the two approaches is that wetland functioning may not necessarily be exclusively dependent on individual species or particular communities, and re-establishment of species may not be a foolproof indicator of re-establishment of the same combination or degree of functioning. Just because ‘it looks like a wetland’ does not automatically translate into ‘therefore it must work like a wetland’. Thus ecosystems whose functioning depends on the progressive accumulation of significant depths of peat or sediment, or relies on their long term biogeochemical dynamics, cannot be created or re-established instantaneously even if the characteristic plants or animals can be successfully introduced and maintained. Time may be a severe limiting factor for the restoration of peatland hydrology for the establishment of carbon stores. It is only now, however, that a more functional approach (see Larson, Chapter 21; Brinson, Chapter 22; Maltby et al., Chapter 23) to restoration objectives is assuming greater significance; and the supporting information and tools are still developing. The multiple values to human societies and the wider environment of the services provided by intact ecosystems are only belatedly gaining recognition, and operational tools for the assessment of ecosystem services are in short supply. We are moving towards a reappraisal of habitats in terms of the actual and potential services they provide, and not the comparatively

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0

Water table depth (m)

0.2 0.4 0.6 0.8 1.0 1.2 1.4 J

F

M

A

M

J

J Month

A

S

O

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Fig. 31.1 Seasonal pattern of water table tolerance zones for a grassland community. Dark grey indicates ‘unacceptable’ conditions of soil moisture, white is ‘tolerable for short periods’ and mid-grey is ‘desirable’ hydrology for the community. (After Wheeler et al. 2004.)

arbitrary target concentrations of chemical constituents or species numbers that have been the benchmarks until now. Methods of assessing the functionality of wetlands are critical needs in this process. Wetland restoration is driven by purpose, and this can be highly diverse. The emphasis in early schemes has often been dominated by the traditional nature conservation criteria – usually the desire to restore a particular habitat or species or, in some cases, simply establish the habitat necessary to support a recovering species, such as the bittern in the UK (see Money et al., Chapter 33). The problem often is the result of the desiccation of the landscape, normally by deliberate actions such as drainage to change land use or its intensity of use. Additionally, it may be due to the unintended consequences of altered water management, such as by dams, surface water and groundwater abstractions, changing flows and lowering water tables. Re-establishment of the ‘good’ hydrological conditions necessary to restore degraded wetlands is often the greater technical challenge. Money et al. (Chapter 33) explore the theoretical and practical basis of achieving hydrological recovery. Not least is the difficulty of not knowing exactly the

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ideal requirements of individual species, though recent work is improving the knowledge base (see Figure 31.1). In this chapter, the authors provide an authoritative guide to the recovery and maintenance of a range of peat-forming wetlands, arguably some of the most difficult to restore. They also deal with systems normally maintained by surface inundation, and those wetland habitats that are neither permanently inundated nor saturated. The water management requirements for restoration are as complex and varied as are the wetlands themselves and the catchments within which they occur. Approaches to the restoration of wetlands for wildlife habitat, dealing with various practical techniques, are examined in Ramseier et al. (Chapter 34). In particular, they emphasise the importance of being able to monitor success. Two options are given. The first are general indicators, such as species diversity or more specific species or groups, which link to particular restoration objectives for example Carex rostrata as an indicator of acrotelm development in peat or dragonflies as an indicator of the ecological health of newly created wetlands (Chovanec and Raab 1997). The second are specific species

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Re-Establishment of Wetland Functioning or groups of species that a project intends to re-introduce to a target area. In this chapter we also have a record of one of the most expensive restoration experiments recorded to date. The transplantation of a mire to make way for the construction of a new runway at Zurich airport (Klotzki and Maltby 1983) was estimated to have cost the equivalent of $125–$416 at 1971 prices (equivalent to around $630–$2100 in 2007) per m2. This translates to $6 300 000–$21 000 000 ha−1, not dissimilar to the cost of high-value development land purchase in many towns and cities (the 2006/7 price of land in the Westminster district of London was nearly £18 million, equivalent to US$35.3 million ha−1). Despite intensive management effort, there were significant changes in species composition, the likely reasons for which are discussed by Ramseier et al. in Chapter 34. The inevitable conclusion is that transplantation is no substitute for conservation or management in situ. Above all, it is an expensive solution, completely out of the question in most situations, and arguably not representing the optimum return for the funding that might be so invested. It is an emergency solution in extremis. Considerable interest focuses on the restoration or creation of wetlands for economic rather then ethical or amenity reasons. Ross and Murkin (Chapter 35) examine the scientific basis and practical management aspects for wildfowl, fur-bearers, fish and crops. An economic basis for restoration can underpin a massive effort in wetland rehabilitation and gain in the extent of the resource. This is underlined by the success of Ducks Unlimited in the USA and Canada. Originally the More Game Birds Foundation, the organisation was formed in the 1930s as a response to the American ‘dust bowl’, when thousands of hectares of wetland habitat were lost during the environmental collapse following the over-exploitation of natural soil resources. Today, Ducks Unlimited has raised many millions of dollars, and has thousands of conservation projects under its belt, which together cover millions of hectares in both the United States and Canada. There is no reason why such targeted economically-motivated objectives related to

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sustainable harvests or hunting cannot also fulfil wider environmental quality objectives such as flood control, water quality and amenity. This strategy of management to achieve multiple benefits is examined in detailed case studies in Section VII. It is the only approach that can secure general support in developing countries and among the world’s poorest communities where management for amenity or traditional conservation values is an ill-affordable luxury, unless it can be linked to activities such as ecotourism capable of enhancing the quality of life without harming that which it seeks to preserve.

DR AIN AGE AN D R EST OR AT ION OF T HE MESOPOT AMIAN MAR SHES It is instructive to consider by way of example one of the modern world’s greatest wetland restoration challenges against the current state of scientific and technical knowledge. The Mesopotamian marshlands are fed by the Tigris and Euphrates rivers, primarily through the flood pulse generated by the annual snow melt from the mountainous headwater regions in Turkey and, to a lesser extent, Iran. Heralded as one of the great ‘cradles of civilisation’, the area is steeped in cultural richness, historical and paleoenvironmental value and ecological diversity. Around 4000 BC, one of the earliest known civilisations was built on knowledge and sustainable use of the Tigris water resources, and gave rise to some of the most ancient writings and accounts of great religious and cultural significance, such as the biblical ‘flood’. The more recent marsh dwellers, marsh Arabs or ‘Madan’, are the link with this cultural past. Their livelihoods are testimony to the intimate relationships between water, life, economy and culture. An analysis of some of the important values of the marshlands are in Table 31.1. These examples indicate: • the significance of the wetlands at local, regional and international scales; • the wider importance of connectivity to other systems especially the marine ecosystems of the Gulf (Maltby 2005).

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Table 31.1 Importance of the Mesopotamian marshlands. (After Maltby 1994.) Unique human community supported by natural resources Productive traditional agriculture Sustainable utilisation of wetland and adjacent land Buffalo and cattle Fishing and birds Reeds and other plants Cultivation Integrated transport Habitat for important populations and species Intercontinental migration Regional biodiversity Rare and endemic species Globally threatened birds, mammals, invertebrates Cyprinid marsh species of high evolutionary significance Linkage to Gulf Hydrological interface between catchment and marine ecosystem Discharge Water quality (nutrients, salinity, contaminants) Sediment Temperature Continuum for movement of economically important fish/ shrimp e.g. Metapenaeus affinis Pomphret Saboor Microclimate Environmental reconstruction Peat/sedimentary deposits

Until recently, the Mesopotamian marshes probably occupied an area of up to 25 000 km2, although estimates vary owing to the technical difficulties of satellite image interpretation, wetland classification and the natural dynamics of flooding (Maltby 1994; Partow 2001). Maltby (1994) recognised three main wetland types – permanent lakes and marsh, seasonal marsh and temporary marsh – in the first study of changes

due to engineering interventions between 1984/5 and 1993. Partow (2001) established an earlier (1973–1976) baseline, and a more recent position in 2000. The United Nations Environment Programme (UNEP) monitoring programme has since extended the time line using both MODIS and Landsat satellite imagery (see http://www. grid.unep.ch/). Some 90% of the Mesopotamian wetlands existing in the 1970s had disappeared by 2000. Most of the loss was from the mid-1980s. Further assessment in 2003 indicated that just 7% of the original (1970s) area remained, restricted almost entirely to a relatively small contiguous area of the Al-Hawiezah marsh and Haur al-Azim straddling the Iraq–Iran border. In 1994, it was estimated that the entire wetland would disappear within 10–20 years (Maltby 1994). In 2003, UNEP estimated that desiccation would be complete within 5 years (see Figure 31.2). The enabling factors and direct causes of wetland loss are now well documented and centre on engineering works, especially dam construction and various deliberate diversions of Euphrates and Tigris flows, preventing the regular annual flood pulse from replenishing the wetlands. The impact of wetland loss was devastating for human lives, quite apart from the effects on environment and ecology. An estimated half-million people were displaced, many killed by the Saddam regime, but because of the loss of the marshes supporting their social and economic livelihood, most of the remainder could be defined as ‘environmental refugees’ (Partow 2001). The loss of the marsh ecosystem raised concerns over the depletion of important bird populations, especially migratory species, and extinction of endemic and rare animal and fish species such as the smoothcoated otter subspecies, the bandicoot rat and the endemic babel Barbus sharpeyi. Increased frequency of dust storms and the effect on regional

Fig. 31.2 The pattern of loss and recovery of the Mesopotamian marshes. In the 1990s, drainage canals had caused the destruction of most of the wetland area that had existed in former times and, by 2002, the tiny fragment remaining was expected to be lost by 2005. Rewetting of the marshlands since then has led to the partial recovery of several areas of the marshlands. Illustration by D. Seger, from UNEP/MODIS data.

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Kuwait

Former extent of wetland

Seasonal wetland

Permanent wetland

1973

Euphrates

Iraq

Ti gr is

Iran

Persian Gulf

2004

Euphrates

gr is

Ti

2000

Iraq

Euphrates

Iraq

Ti gr is

Kuwait

Kuwait

Iran

Iran

Persian Gulf

Persian Gulf

2005

Euphrates

Iraq

Ti gr is

2002

Euphrates

Iraq

Ti gr is

Kuwait

Kuwait

Iran

Iran

Persian Gulf

Persian Gulf

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climate and loss of fish nursery grounds are other examples of impact, which extend well beyond the boundaries of the freshwater wetlands (Maltby 2005). The uncoordinated restoration of the marshes began in 2003 almost immediately after the fall of the Saddam regime. Communities took the initiative to breach the structures that had been placed to prevent flooding. More widespread and controlled inundation followed with governmental support. In Spring 2005 UNEP reported recovery of over 50% of the former marshland extend present in the 1970s (Figure 31.2; see also http://inos.grid.unep.ch/). Field surveys have also documented a surprising rate and degree of re-establishment of many species (Richardson and Hussain 2006; Abed 2007). The new government of Iraq has recognised the importance of marshland restoration as an icon to some, the mainstay of desired livelihoods to others, but for yet others an impediment to economic development and improved human wellbeing. Internationally, there is strong interest in restoration to re-establish biodiversity and to maintain the environmental and ecological quality of the Gulf, thus safeguarding valuable economic resources, however, there are also many competing demands on the water essential for restoration from neighbouring countries and within Iraq itself. The overarching questions commonly raised concerning the restoration of the Mesopotamian marshes are: 1 What do the local communities, and especially the Marsh Arabs, actually want? 2 What proportion of the former wetland can and should be recovered? 3 What are the constraints to achieving agreed objectives in the area? It is still not clear what is the agreed common vision for the marshes, notwithstanding the establishment of an Interministerial Centre for the Restoration of the Iraqi Marshes (CRIM) and a growing number of more or less comprehensive technical reports and assessments (e.g. Iraq Marshlands Restoration Programme Action Plan 2004; New Eden Master Plan for Integrated Water

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Resources Management in the Marshland Area 2006). Conversations between the present author and both individual Madan as well as tribal leaders have highlighted the potential dilemmas. Whilst wishing to retain traditional culture and the contentment stemming from the wetland surroundings, the Madan also want the benefits of economic development: clean water, electricity, access to schools, clinics, markets and other features that enhance human well-being. This is a not unfamiliar story from throughout the developing world. The message for the decision-makers is not a straightforward one; restoration needs to be part of a comprehensive and highly integrated sustainable development strategy. This will need to take full account of many, and sometimes widely separated, stakeholders representing different spheres of interest (Figure 31.3). There is no doubt at all that marsh restoration is technically feasible; it has been progressing more or less naturally already over the last 5 years. Questions remain regarding the optimum extent and specific location of effort and the hydrological and management regime that can yield the greatest benefit to local people as well as the Iraqi population as a whole, and indeed the wider international community. The answers will rest substantially in decisions on the best use of water resources upstream as well as the outcome of bilateral or multilateral discussions with neighbouring countries, especially Turkey and Iran, who are drawing down water that otherwise could support restoration. At some stage it will be essential to establish a river basin-scale dialogue among countries and various sectoral interests (especially agriculture, power and related industry, fisheries, conservation and regional development) to agree a basis for the optimum and shared use of water resources in the Tigris and Euphrates. The ability of the restored marshes to function and contribute to a wide range of human benefits, often not always realised or quantified, should be a key part of these discussions (Table 31.2). Re-flooding alone does not constitute restoration. The estimation of the optimum extent of rehydration must take account of the dynamics (e.g. through-flow), connectivity, and associated

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Re-Establishment of Wetland Functioning Spheres of interest International actors

River basin management Scientific and technical support Capacity building and training

Regional interests National

Local people

e.g. oil other development priorities upstream activities

Stakeholders

e.g. Water quality/regime fisheries biodiversity

Quality of life: • Social, economic, cultural • Health concerns • Sustainable environment

Fig. 31.3 Different sectoral groups with different interests and objectives operate at different scales to influence the outcome of restoration objectives in the Mesopotamian marshes. (Maltby 2003, presentation to the UNEP Round Table on Iraq, Geneva.) Table 31.2 The causes of wetland degradation, their consequences, and options for mitigation. (From Maltby 2003.) Potential issue

Consequences

Restoration/mitigation options

Fisheries decline

Economic loss Ecosystem change

Rehabilitate marshlands Develop alternative fish stocks

Contamination of fish/shrimp

Human health impact Economic loss Food web disruption

Re-establish ‘natural’ filters in delta/coastal margins

Degradation of water quality

Fisheries impacted Ecosystem change

Reduce use of fertilisers, pesticides in catchment, trap other contaminants

Reduced freshwater inflow

Salinity increase Temperature regime Productivity Ecosystem change

Increase releases from upstream structures

Influx toxins/contaminants

Human health Fisheries impact Food web disruption Ecosystem change

Re-establish ‘natural’ marshland filters; reduce contaminant releases to water by traps/filters etc.

Reduced sediment influx

Coastal erosion Fisheries decline

Remove engineering structures/transport sediment to by-pass structures

Increased sediment influx

Increased turbidity Impact on coral reefs

Sediment retention structures

Biodiversity loss

Ecosystem instability

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landscape and socio-economic characteristics that will realise the desired benefits for society. None of this is constraint free. Restoration of the Mesopotamian marshes needs to compete with the priorities of government time and investment. The problems of security, urban, water and power supplies and health currently attract greater attention. Progressive drainage of the Haur al Azim wetlands on the Iranian side of the Howeizeh marshes, and completion of a dyke artificially separating these parts of the same contiguous ecosystem, may threaten the integrity of the restoration effort. Major, currently untapped oil reserves, coincide with the former marshland area. Economic factors will significantly influence actions to re-flood a large area of landscape.

R E F E R E N CE S Abed 2007. Status of water birds in restored southern Iraqi marshes. Marsh Bulletin 2(1), 64–79. Chovanec A. and Raab R. 1997. Dragonflies (Insecta, Odonata) and the ecological status of newly created wetlands: examples for long-term bioindication programmes. Limnologica 27(3–4), 381–392.

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Klotzki F. and Maltby E. 1983. Mires on the move in Europe. Geographical Magazine LV, 346–349. Maltby E. (editor) 1994. An Environmental and Ecological Study of the Marshlands of Mesopotamia. Amar Appeal Trust, London. Maltby E. 2003. Mesopotamian Marshlands. Opportunities for Marshland Restoration. Unpublished presentation to UNEP Roundtable, 23 May 2003, Geneva. Maltby E. 2005. Approaches to the re-establishment of the freshwater-marine continuum through the Mesopotamian Marshes: regional, sectoral and transboundary challenges. Technical Presentation. High-level meeting on the restoration of the Mesopotamian Marshlands and the marine environment. Bahrain, February 2005. Regional Organisation for the Protection of the Marine Environment, United Nations Environment Programme. Partow H. 2001. The Mesopotamian Marshlands: Demise of an Ecosystem. Report, United Nations Environmental Program, Geneva, 58 pp. Richardson C.J. and Hussain N.A. 2006. Restoring the Garden of Eden. An ecological assessment of the marshes of Iraq. Bioscience 56(6), 477–489. Wheeler B.D., Gowing D.J.G., Shaw S.C., Mountford J.O. and Money R.P. 2004. Ecohydrological Guidelines for Lowland Wetland Plant Communities. Environment Agency (Anglian Region), Peterborough.

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32

Restoration of Wetland Environments: Lessons and Successes A R NOL D G . VAN D ER VALK Department of Ecology, Evolution and Organismal Biology, Iowa State University of Science and Technology, Ames, USA

IN T R O D U CT ION Ecological restoration is the repair of damaged ecosystems, the rebuilding of destroyed ecosystems or the replacement of one ecosystem by another (Jackson et al. 1995). Typically, the focus of ecological restoration is the repair or building of indigenous, natural ecosystems, for example the floodplain wetlands of the Kissimmee River (Toth 1995), but it can also be the repair of manmade ecosystems (e.g. floating fens that developed in peat excavations in the Netherlands (van der Valk and Verhoeven 1988; Money et al., Chapter 33). One of the most important developments in wetland ecology in the last 25 years has been the increasing interest in the restoration of wetlands, particularly in the United States and Europe (Kusler and Kentula 1990; Wheeler et al. 1995; Hey and Philippi 1999; Pfadenhauer and Grootjans 1999; Klotzli and Grootjans 2001; Gorham and Rocheford 2003; Money et al., Chapter 33) but also in many other parts of the world (Streever 1999; Murphy and Munewar 1999; van Andel and Aronson 2006). Although most wetland restoration projects are still relatively small, a few hectares or less, recent and proposed projects like the ongoing Kissimmee River

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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(Dahm 1995) and The Everglades restoration projects (Davis and Ogden 1994) in Florida, USA, aim to restore tens of thousands of hectares of wetland. As the number and scale of projects continue to grow, an evaluation of what we know and what we don’t know about the ecological restoration of wetlands is pertinent in order to improve the planning, design and implementation of future projects. So far, little established ecological theory seems to have been utilised in the design or implementation of wetland restoration projects (van der Valk 1998; Cole 1999). As Pickett and Parker (1994) have pointed out, ecological restoration has been practised for a long time, but it has only recently begun to develop as a separate discipline with its own theories. Are these new restoration theories original and useful? To evaluate them, they need to be scrutinised to determine if they provide any new insights into restoring wetlands. The extensive literature, both descriptive and theoretical, on vegetation and ecosystem development (e.g. McIntosh 1985; Glenn-Lewin et al. 1992; Golley 1993; van Andel et al. 1993; Christensen et al. 1996; Samuels and Drake 1997; Gurevitch et al. 2002; Falk et al. 2006; van Andel and Aronson 2006) provides a gauge with which to appraise these restoration theories. This chapter reviews lessons that have been learned about the restoration of wetlands, issues that remain to be resolved and suggestions for improving future projects. Two theoretical issues are examined first: (i) the relation of the new

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ecological restoration theories to established theories of succession; and (ii) the similarity of created or restored wetlands to natural wetlands. The remainder of the chapter is an overview of major practical issues faced when planning a project. Before we consider these theoretical and practical issues, however, some basic terminology is presented. Terminology The literature on ecological restoration contains a bewildering number of terms to describe various kinds of ecosystem repair and building projects. Terms like enhancement, rehabilitation, replacement, reclamation, restoration, creation, mitigation, management, and so on, are all in use and are often used interchangeably. The plethora of terminology, in part, reflects the fact that ecological restoration projects can vary greatly in both initial conditions at a site and desired outcomes. In this chapter, a slightly modified version of the terminology proposed by Lewis (1990) will be used. To simplify matters only two broad classes of projects will be considered: restorations and creations. Restorations are projects in which a wetland is re-established on a site where a wetland previously existed. The re-establishment of a prairie pothole by breaking or plugging drainage tiles that had been used to drain the wetland 50 years ago is an example. The restored wetland need not be identical or even similar to the wetland that existed on the site previously, although it could be. Wetland creations are projects in which a wetland is established on a site where previously a wetland had not existed. For example, a shallow excavated basin that is planted with wetland species in an area that was formerly upland. Ecological restoration projects are always done for a purpose. Many are done to replicate as closely as possible a wetland type that previously existed on the site or still exists at another location nearby (historic restorations or creations). The main criterion for success of historical projects is the demonstration that the plant

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and animal species that characterise this wetland type are present. Most other projects are done to establish a wetland that has certain functional or aesthetic characteristics (functional restorations or creations), for example a wetland that will act as a sink for pollutants in agricultural runoff. The main criterion for success of functional projects is the demonstration that the wetland does function as expected, for example 90% of the phosphorus entering the wetland is retained. Many wetland creation and restoration projects, particularly in the United States, are mitigation projects. Mitigation is the required replacement of wetlands lost due to development. The intent is to replace the values of the lost wetland to society by creating or restoring a comparable wetland. In the United States, mitigation is usually required in order to obtain the permits needed to fill existing wetlands (Salvesen 1994; Committee on Mitigating Wetland Losses 2001). The re-introduction of animal species into existing wetlands, for example the re-introduction of whooping cranes or mute swans in areas where they were extirpated, is a re-establishment project. Re-establishment projects are outside the scope of this chapter, but are covered in two other chapters by Ramseier et al., Chapter 34 and Ross and Murkin, Chapter 35. They are highly species specific and often do not involve either wetland creation or restoration per se. From an ecological perspective, both creation and restoration projects can be divided into three stages: environmental, biological and ecosystem or functional restoration. The relative importance of these can vary significantly among projects. Designing and implementing the first two stages, which establish appropriate environmental conditions and plant and animal communities, are the chief responsibility of the restorationist (Money et al., Chapter 33). Although it may be possible to accelerate the development of ecosystem restoration, for example by adding nutrients or organic matter to the soil, it is impossible for the restorationist to reinstate most ecological functions of a wetland directly. Many ecological functions, such as secondary production, are dependent on the growth and spread of

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Restoration of Wetland Environments plant species and the build up of a fallen litter layer. In other words, they are a consequence of the accumulation of energy and materials (Odum 1969, 1971; van der Valk 1985; Golley 1993).

T HE OR E T ICAL IS S UE S Ecological restoration theories There are many ecological concepts and theories that are potentially relevant to the planning, implementation and evaluation of ecological restoration projects, including niche, competition, disturbance, island biogeography and succession (Palmer et al. 1997; Cole 1999; Zedler 2000a; Davis and Slobotkin 2004; Winterhalter et al. 2004; Falk et al. 2006; van Andel and Aronson 2006). Of these, succession arguably has the most direct relevance to ecological restorationists. Human intervention to remove impediments to the establishment of vegetation is what ultimately distinguishes creation and restoration from natural succession (Figure 32.1). Human intervention can reduce or eliminate environmental barriers to plant and animal establishment (e.g. low substrate fertility), and can also eliminate most of the uncertainties associated with dispersal, establishment and growth of plant species. In short, creation and restoration are simply accelerated succession. Because the development of suitable wetland vegetation or a suitable wetland ecosystem is their ultimate goal, wetland restorationists have begun to theorise about the development of restored and created wetlands. To date, two theories (self-design and designer) have been proposed that purport to describe the development of created and restored wetland ecosystems (Mitsch and Wilson 1996; van der Valk 1998; Mitsch et al. 1998; Middleton 1999). These two theories are described and then examined in light of the literature on succession. Self-design According to Mitsch et al. (1998), ‘Self-design relies on the self-organisation ability of

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ecosystems, in which natural processes contribute to species introduction and selection. In selfdesign, the presence and survival of species due to their continuous introduction and that of their propagules (i.e., many species are introduced but few are chosen) is the essence of the successional and functional development of an ecosystem’. They go on to say that ‘[i] n the context of ecosystem restoration and creation, self-design means that if an ecosystem is open to allow “seeding,” through human or natural means, of enough species’ propagules, the system itself will optimise its design by selecting for the assemblage of plants, microbes, and animals that is best adapted for existing conditions’. The self-design theory postulates that the development of a created or restored wetland is a deterministic process whose endpoint is prescribed by environmental conditions. The environment will allow only some species to become established, and the characteristics and functions of the resulting ecosystem are the product of its environment. Because environmental conditions are the prime determinant of a restored ecosystem’s composition, structure and function, all ecological restoration projects with similar environmental conditions will converge toward the same ecosystem. In short, self-design is an equilibrium theory that postulates that self-organisation directs the development of an ecosystem to a fixed endpoint determined by environmental conditions. Self-design theory implies that ecological restoration of wetlands is simple and easy. According to Mitsch et al. (1998), all that is needed is the ‘seeding’ of the restoration or creation project by nature or humans. How to decide whether nature or humans should do the seeding is not specified. Another implication of self-design is that restoration and creation projects cannot go wrong. In Mitsch’s words, ‘... the system itself will optimise its design by selecting for the assemblage of plants, microbes, and animals that is best adapted for existing conditions’. Hence, the primary emphasis in wetland creation and restoration projects should be engineering the correct hydrology and other environmental conditions. Self-organisation will take care of the rest; that is, nature will do

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(a) Seed dispersal

Seeds

Succession

Seedlings

Maturation

Adults

f(G)

f(E) F(A), f(D), f(H), f(P)

f(C), f(D), f(H), f(P)

f(A), F(C), F(D), F(H), f(P)

Litter

A – Age C – Competition D – Disturbance E – Environmental conditions G – Growth H – Herbivory P – Pathogens (b) Sowing of seeds

Seeds

Planting of seedlings

Succession

Seedlings

Maturation

Adults

f(G)

f(E) F(A), f(D), f(H), f(P)

Planting of adults

f(C), f(D), f(H), f(P)

f(A), F(C), F(D), F(H), f(P)

Litter

A – Age C – Competition D – Disturbance E – Environmental conditions G – Growth H – Herbivory P – Pathogens Fig. 32.1 Simplified succession (a) and ecological restoration (b) models with all potential feedback loops and interactions ignored. Human interventions that could alter plant–herbivore, plant–pathogen and plant–plant interactions during restoration are also not shown. (From van der Valk 1998.)

the remainder of the ‘engineering’. In other words, wetland creation and restoration requires understanding of hydrology but not much knowledge of wetland ecology (especially of topics like plant dispersal, establishment and growth). Designer theories Designer theories postulate that various kinds of ecosystems can be established and will persist

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in a given ecological restoration project. Because the wetland ecosystem that develops is, to a large extent, a function of the plants that become established, it is possible to establish a variety of ecosystems. Designer theories differ from self-design theories in that there are no fixed endpoints. In other words, there is no convergence towards a specific ecosystem under a given set of environmental conditions. There are, however, limits on the kinds of species that can be established and

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Restoration of Wetland Environments that can persist in a given environment. Mitsch et al. (1998) describe designer theories as ‘botanical engineering’. Although most restorationists do not explicitly acknowledge it, designer theories provide the theoretical framework in which they are working. One of the major benefits of designer theories is that they focus the restorationists’ thinking on the ultimate ecological realities of establishment, growth and extirpation of species (Garbisch 1990; Galatowitsch and van der Valk 1994; van der Valk 1998). Designer theories require the restorationist to select suitable species to establish; to decide where, when and how to establish them; and how best to manage the site to ensure their survival. In short, all available information about desirable plant and animal species in the ecological literature can and should be used to improve the design and implementation of an ecological restoration project. Succession theories Are these ecological restoration theories new? Do they have any antecedents in the literature on succession? The literature on succession is enormous, and it is impossible to summarise even the highlights in this chapter. There are, however, excellent summaries of this literature in Miles (1979), McIntosh (1985), Glenn-Lewin et al. (1992) and van Andel et al. (1993). The succession literature as it relates specifically to wetlands has been reviewed by van der Valk (1981, 1982, 1985, 1987, 1992, 2006), Keddy (1999, 2000) and Middleton (1999). In fact, there are only two basic types of succession theory, although there are many variations on these two themes: equilibrium theories and non-equilibrium theories. Both types will be briefly outlined and then compared with the self-design and designer theories being put forward by restorationists. Formal equilibrium theories of succession originated with Clements (1916, 1928) who somewhat cryptically defined succession as the development of a climatic climax formation, that is, a self-perpetuating vegetation type in equilibrium with the regional macroclimate. Fifty years

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later, Odum (1969), who was interested in the development of ecosystems, altered Clements’ model of succession from the development of a self-perpetuating plant community to the development of a self-perpetuating ecosystem. Like Clements, many early ecosystem ecologists viewed ecosystem development as a deterministic phenomenon that is driven to a fixed endpoint by environmental conditions (O’Neill 2001). The first formal non-equilibrium theory of succession was put forward by Gleason (1917, 1926, 1927). According to Gleason, succession is simply a change in the composition of the vegetation that is the result of the migration of populations into or out of an area. Each population has its own unique or individualistic distribution that is a result of both its migrational history and its environmental tolerances. Gleason saw the landscape as being covered with a constantly changing cover of vegetation and considered vegetation change as a population-level phenomenon. The composition of the vegetation of an area was simply the result of overlapping populations, not the result of some community-level or ecosystemlevel phenomenon that directs the assemblage of a fixed group of species. Divergence, not convergence, is the key feature in Gleason’s thinking. Plant ecologists have spent decades arguing about succession, particularly the relative merits of the Clementsian and Gleasonian views of succession (McIntosh 1985). Almost all of the studies of succession in wetlands and other vegetation types suggest that the non-equilibrium models of changes in species composition (Gleasonian succession) fit the data much better than the equilibrium models (McIntosh 1985; Glenn-Lewin et al. 1992). Some ecologists, however, believe that it may be possible to develop a model of plant community dynamics that encompasses elements of both Gleason’s individualistic/non-equilibrium and Clements’ organismal/equilibrium models of succession (Moore 2001; Lortie et al. 2004). Nevertheless, the major conclusions from reviews of successional studies are that (i) ‘[v] egetation dynamics are based on the population processes of the constituent plants’. (Glenn-Lewin et al. 1992);

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(ii) succession is usually a non-equilibrium process (Sprugel 1991); and (iii) convergence toward a fixed endpoint under the same environmental conditions does not always occur (Samuels and Drake 1997). Sprugel (1991) succinctly summarised the modern consensus: ‘... there are often several communities that could be the “natural” vegetation for any given site at any given time’. Many studies of ecosystem development have demonstrated that it is often an orderly and predictable process with profound consequences for the structure and functioning of ecosystems (Peet 1992). Samuels and Drake (1997), who reviewed the literature on convergence, suggest that ‘[m] any studies reporting convergence found convergence at rather crude structural levels (e.g. guild representation) while those reporting divergence observed divergence at finer levels of scale like species composition’. Most functional characteristics of ecosystems are simply the result of the growth of plants and the fate of this primary production. It is universally recognised that various ecosystem functions change as constituent species grow and ecosystems develop (van der Valk 1985; Peet 1992; Golley 1993; Christensen et al. 1996; O’Neill 2001; Gurevitch et al. 2002). A variety of non-equilibrium models of ecosystem development have been developed to describe and predict these functional changes (see Urban and Shugart 1992) that do not invoke self-design or self-organisation. Functional convergence does not always occur, although it is more likely than compositional convergence during succession. Restoration versus succession theories A comparison of the self-design and designer restoration theories indicates that they are essentially re-statements of equilibrium and nonequilibrium succession theories respectively. Restoration ecologists have not provided any new insights into the development of vegetation or ecosystems. They are simply recycling old theories. Rather than re-fight the ecological succession wars of the 1950s and 1960s (McIntosh 1985), restorationists would be well advised to

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recognise that the development of a created or restored wetland is a non-equilibrium process that can best be understood and described using non-equilibrium succession theory (Pickett and Parker 1994; van der Valk 1998; Middleton 1999; Davis and Slobotkin 2004). Not only is the self-design theory inadequate to deal with the conceptual complexities of created or restored wetland ecosystem development, its adoption also threatens to undermine sound ecological restoration practice. Mitsch et al. (1998) complained about American mitigation policies that call for the ‘... survival of selected organisms ...’ because they result in wetlands that are not self-designed and hence by definition are not sustainable. They provide no evidence that designed wetland ecosystems are, in fact, unsustainable. They also fail to recognise that historic ecological restoration projects are fundamentally different from functional projects. In the former case, sustainability, which seems to imply no human intervention at all, is not the goal. Many historic creations and restorations require management for their long-term survival, as a substitute for a variety of disturbance and other factors that no longer occur naturally. This has long been recognised by restorationists working on both wetland and terrestrial systems. For example, The Tallgrass Restoration Handbook (Packard and Mutel 1997) devotes as many chapters to post-establishment management and monitoring as it does to planting techniques. If self-design prevails and active intervention in establishment and management is eliminated, most wetland creation and restoration projects will come to be dominated by a handful of invasive wetland species. This is often the case today for projects that do not involve the active reestablishment of species (Galatowitsch and van der Valk 1994, 1996; Seabloom and van der Valk 2003a,b; Ramseier et al., Chapter 34). Because the self-design theory downplays the importance of detailed planning for the biological restoration stage and the need for post-restoration monitoring of creation and restoration projects, its adoption would be disastrous for projects whose major goal is to preserve or enhance biological or

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Restoration of Wetland Environments landscape diversity. Inevitably, this will result in increased failures of these projects that could jeopardise funding for future projects. Some restorationists have argued that for projects where the probability of revegetation is high because of natural dispersal, for example riverine wetland restoration projects, or for projects in which seeds of a variety of wetland species are sown, that the self-design and designer theories are of little practical significance because they will result in development of the same kind of vegetation. Other restorationists have suggested that both theories are largely irrelevant because the outcomes of most projects are largely unpredictable, for example, Zedler (1999, 2000a). The bottom line is that self-design is inappropriate for historic creation and restoration projects. For functional projects, for example, wetlands constructed for sediment removal, assuming seed dispersal to the site is not a problem, un-assisted re-vegetation of a restore or created wetland will result in an acceptable wetland mainly because the precise composition of the vegetation is irrelevant to the goals of the project. Created and restored versus natural wetlands Can wetlands be created or restored that are identical to natural wetlands? This is perhaps the most fundamental question facing wetland restorationists. The outcome of any project is a function of many factors, including the initial state of the site, the amount of effort put into it and the measures of success used to evaluate the project (Zedler 1999). Chance undoubtedly plays a major role in the outcome of many projects. Regardless of the amount of time and effort that went into it, no created or restored wetland will ever duplicate exactly a lost or extant natural wetland. There are many reasons why newly created or restored wetlands will not be comparable to wetlands that existed historically or exist currently in the same region. Some wetland species or ecotypes may have been extirpated locally or, much more rarely, have gone extinct. New wetland species may have become established in the area. Regional changes in water chemistry, especially nutrient levels, are

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common. Regional changes in hydrology, due to lowering or raising of water tables or changing precipitation patterns, might have occurred. Such changes make it impossible to establish wetlands that are identical to those that existed previously. This seems to suggest that it is never possible to have successful historic restoration projects. This line of reasoning, however, is based on a fundamental misconception of the static nature of natural ecosystems, that is, an equilibrium concept of ecosystem development. The goal of historic ecological restoration projects should not be to establish wetlands that are identical to those of the past, which is impossible, but to preserve as much as feasible of the biodiversity and genetic diversity of those wetlands. A more realistic question is how similar to natural wetlands can created and restored wetlands become? Wetlands are often created or restored that, at least superficially, resemble natural wetlands (see Figures 32.2 and 32.3.) An operational restatement of this question is ‘were the goals of the project met?’ This is the only meaningful criterion for judging the success or failure of a project. The reliability of this judgement is a function of the adequacy of the performance measures selected and of the project monitoring done (Kusler and Kentula 1990; Reinartz and Warne 1993; Galatowitsch and van der Valk 1996; Brown 1999; Moore et al. 1999). By some, success would be claimed for a created or restored wetland if a single duck landed on it once. For others, it would be claimed only after five or more years of rigorous quantitative sampling had shown that the created or restored wetland’s species composition and functions were comparable to those of reference wetlands. Many projects are simply declared to be successes without any monitoring data collected to see if they have met their stated objectives. In fact, relatively few wetland projects or groups of projects have been evaluated in detail. When this has been done, many created and restored wetlands are judged to be failures because they do not meet their original objectives (Malakoff 1998). Gwin et al. (1999), Shaffer et al. (1999), Shaffer and Ernst (1999) and Magee et al. (1999)

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Fig. 32.2 Final stages (June 1992) of the construction of a sedge meadow swale at the Des Plaines Wetland Demonstration Project north of Chicago, Illinois, USA (photograph Joy Marburger). See Plate 32.2 for colour version of this image.

Fig. 32.3 Two-year-old (May 1994), created sedge meadow swale at the Des Plaines Wet land Demonstration Project north of Chicago, Illinois, USA (photograph Joy Marburger). See Plate 32.3 for colour version of this image.

examined wetlands constructed for mitigation in the Portland, Oregon, USA, metropolitan area. Their studies indicate that wetlands were created that had no natural analogues in the area, that had very low soil organic matter content, that often had plant communities dominated by exotic plant species, and that had a low diversity of native wetland plant species. These results and other similar studies suggest that the practice of wetland creation and restoration leaves much to be desired (Salvesen 1994). Most of these failures

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seem to be due to (i) poor design work, (ii) inadequate supervision of projects by regulatory agencies and (iii) construction firms cutting corners to save money. Poor project design usually results in inappropriate basin topography, atypical hydrology and unsuitable plant communities. Even with projects that are well designed and executed, the created or restored wetlands may not closely resemble their natural counterparts. Zedler and Callaway (1999) examined long-term data from Sweetwater Marsh, a created coastal

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wetland in Southern California, and from other constructed wetland sites, to see if properties and functions of created and restored wetland ecosystems followed the trajectories predicted by equilibrium models (Figure 32.4). Sweetwater Marsh was constructed to create habitat for several endangered species, including the light-footed clapper rail (Rallus longirostris levipes). One of the major problems with this restored wetland was that the grass that was established did not grow as tall as expected due to the low nutrient status of the coarse, sandy substrate (Zedler 1993; Gibson et al. 1994). Because of the low stature of the vegetation, the light-footed clapper rail was not attracted to the site (Malakoff 1998). Adding nutrients to improve the growth of the grass, however, resulted in the growth of an unwanted species that out-competed the grass. Zedler and Callaway (1999) concluded that functional

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Fig. 32.4 Soil nitrogen and organic matter and Spartina foliosa stem length and shoot density in created and nearby natural wetlands in San Diego Bay, California, USA (Zedler and Callaway 1999). (a) Total Kjeldahl nitrogen (TKN) in surface soils; (b) Soil organic matter content (loss on ignition); (c) Spartina foliosa total stem length at the end of the growing season; and (d) Number of Spartina foliosa shoots taller than 90 cm.

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Restoration of Wetland Environments

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development does not always follow the simple trajectory models that are widespread in the restoration ecology literature. Instead, there is often high inter-annual variation and lack of directionality to measures of functional development both in constructed and natural wetlands (Figure 32.4), that is both are non-equilibrium systems (see also Simenstad and Thom 1996). Others, however, have found that, in some restored or created wetlands, there are predictable trajectories for some performance measures (Craft et al. 2003). As noted previously, ecologists now generally agree that succession is a non-equilibrium process that does not lead to one fixed endpoint. From an initial set of environmental conditions (Seabloom et al. 1998) and for a variety of reasons (Table 32.1) many different ecosystems can develop during succession and as a result of ecological restoration that vary in composition and

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Table 32.1 Potential sources of variation in community and ecosystem development (Samuels and Drake 1997). Source of variation

Examples

Environmental gradients Sequence events

Altitude, moisture, nutrient Community composition depends on order species arrive Development highly dependent on initial conditions Alternate endpoints possible under same conditions Variation caused by environmental stochasticity or species traits

Chaotic dynamics Indeterminate trajectories Random noise

function. What should be the ‘target’ for an ecological restoration project? How can the success of such a project be judged? The uncertainties associated with the development of ecosystems would seem to make ecological restoration a process whose outcome is ultimately unknowable. As noted previously, this theoretical dilemma is the result of viewing succession and ecological restoration as discrete events. In reality, all ecosystems are constantly changing. They are dynamic systems that are perpetually adapting to short and long-term changes in environmental conditions and to various kinds of disturbances. Because it has no fixed endpoint, succession can be directed by both natural forces and human manipulation (Luken 1990). Ecological restorations can also be directed toward desired endpoints, but this intervention must continue in perpetuity. To keep a wetland developing along a specific trajectory will require periodic interventions. What those interventions should be, and when they need to be made, requires a sound understanding of the factors that affect the succession of comparable natural wetland ecosystems. Common interventions include manipulations of water levels, fire, grazing, fertilisation and various weed control methodologies. Although a created or restored wetland will never be identical to a given natural wetland,

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it is, nevertheless, possible to create or restore wetlands that will closely resemble natural wetlands in function and composition. To do this, however, requires an understanding of the ecology of wetlands, careful planning and execution of projects, and an adaptive management strategy to keep the wetland on its desired developmental trajectory.

PR ACT ICAL ISSUES There are many issues that need to be considered when planning ecological restoration projects. These vary depending on the type of project (creation vs. restoration); the objectives of the project (functional vs. historic); the size and scale of the project; the legal framework in which the project sits; the location of the project; the training and experience of the restorationists and the amount of money available. This section is not designed to be a handbook on planning and carrying out ecological restorations or a compendium of engineering guidelines for constructing dykes, water control structures or sizing pumps. Instead, it is designed to provide a brief overview of the most important ecological issues that should be considered by restorationists when planning projects and some guidelines for dealing with them. A typical management process for a restoration project involves a series of stages (Figure 32.5) that begins with the setting of project goals and performance measures and ends with an assessment of whether these goals were met or not. Only those stages of a project that involve its planning and assessment will be discussed. Project planning Ecological restorations have to be carefully planned. The amount of planning required varies with the scope and novelty of the project. Who should be involved in the planning also depends on the scope, novelty and visibility of the project. Planning for the restoration of the 9508th prairie pothole to be restored in the Upper Midwest in

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Specific project goals and performance measures

Site selection

Site evaluation (identification of constraints)

Restoration plan

Environmental restoration Project assessment

Monitoring Biological restoration Project management Adaptive management

Fig. 32.5 Typical stages in the design, implementation and assessment of ecological restoration/creation projects.

the USA can be minimal. This is a routine smallscale project that will require little engineering and little money. On the other hand, the restoration of the floodplain of the Kissimmee River in Florida, for which there was no precedent, required years of planning (Dahm 1995). Planning of the Kissimmee restoration involved large numbers of people including local land owners, engineers, ecologists, hydrologists and lawyers, as well as local, state, and national politicians, government agencies and non-government agencies. The main goal of planning is to reduce uncertainties and thereby to increase the probability of the project being successful. Another major goal of planning is to get or retain public and political support for the project. Most restoration projects are the result of public policy decisions resulting from public pressure, and many projects are funded with public money. Consequently, societal values need to be considered in the planning and assessment of projects. If the project has been ‘sold’ because it will improve water quality, some way to assess the project’s effect on water quality needs to be included in the performance measures selected

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and in the project’s assessment. There is a need to involve the public in the planning process not only to retain their support, but also to educate them about the underlying science that will be used to design the project. The latter can be essential because scientific perceptions of the goals of a restoration project are sometimes at variance with public perceptions of what the goals should be (Higgs 2005). Potential real or perceived negative consequences of a project should also be considered and addressed during the planning phase, for example, an increase in mosquitoes and mosquito-born diseases (Willott 2004). One essential aspect of the planning process is the identification of uncertainties and risks. Where might problems develop because of a lack of knowledge about the wetland type being restored or created? For most projects, there may be considerable uncertainty about the actual hydrology of the new wetland, the best way to establish desirable plant species and the potential invasion of the site by unwanted weed species. The project plan needs to identify when problems due to uncertainties may arise; it should contain provisions for monitoring the project to detect

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problems that develop; and it should consider possible strategies for correcting them, if necessary, through adaptive management (Figure 32.5). Numerous guidelines and suggestions for planning ecological restorations have been proposed (e.g. Garbisch 1990; Kentula et al. 1993; Galatowitsch and van der Valk 1994; White and Walker 1997; Weinstein et al. 1997; Galatowitsch et al. 1998; Geist and Galatowitsch 1999; Lake 2001; Schreiber et al. 2004; Ramseier et al., Chapter 34). There are also many sources of information about how to create or restore specific wetland types (e.g. Broome 1990; Galatowitsch and van der Valk 1994; Getter et al. 1984; Kadlec and Knight 1996; Lewis 1982; Wheeler et al. 1995; Wilcox and Whillans 1999; Zedler 2000b; Money et al., Chapter 33, and Ramseier et al., Chapter 34). Middleton (1999) provides a comprehensive overview of ecological restoration and management for freshwater wetlands and Zedler (2000b) for tidal wetlands. The Committee on the Restoration of Aquatic Ecosystems (1992) compiled a restoration project checklist that covers project planning and design, project monitoring and post-restoration assessment (Table 32.2). Project goals and performance measures The first and most important step in the planning process is to establish a consensus about the scope and nature of the project and about its societal, legal and scientific goals and to develop performance measures that reflect the project’s goals and that can be used to decide if these goals were achieved (Kentula 2000). Without articulated goals, it is impossible to design a suitable project. Two types of project goals are often needed, general goals and specific goals. The general goals should reflect the perceived value of the project by the general public and policy makers, for example, to improve water quality or to increase duck production. They should be written in language that is readily understood by non-scientists. The specific goals are the quantifiable goals

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of the project, for example, area to be restored or type of vegetation to be established. Specific goals are often based on the characteristics of former or extant wetlands of the type being restored, often called reference or target wetlands. Careful attention also needs to be given to the resources (expertise, labour, equipment, etc.) that will be needed to carry out the project. The restoration plan that is developed should incorporate any short- and long-term management that will be needed to direct the development of the new wetland ecosystem toward the project’s goals or objectives. For a large project, a time line may be needed which describes when certain specific goals are expected to be met either completely or partially. Without well-defined specific goals, it is impossible to develop meaningful performance measures, to determine whether a project was successful or not, or to determine whether there is a need to intervene in the wetland’s development before the environmental and biotic restoration phases have been completed (Figure 32.5). Performance measures (also called success criteria and expectations) are physical, chemical or biological attributes of the project that will be monitored during or, at a minimum, at the end of a project. For example, a performance measure could be the percent of a restored wetland covered with a certain species or vegetation type. They obviously need to be linked to the project’s general and specific goals. Some performance measures, however, may be needed to meet legal or permit requirements. They need to be appropriate given available personnel, equipment, time, money and so on available for monitoring. Toth and Anderson (1998), Weinstein et al. (1997) and Ruiz-Jaen and Aide (2005) provide overviews of the types of performance measures or expectations that need to be developed to evaluate a project and how to use these measures to trigger adaptive management of the project. Performance measures can include species diversity, presence of target plant or animal species, species abundance (cover, density), species biomass, soil conditions (nutrients, texture, organic matter), nitrogen fixation rate, flood

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Table 32.2 Restoration project planning guidelines (Committee on the Restoration of Aquatic Ecosystems (1992)). Project planning and design • Has the problem requiring treatment been clearly understood and defined? • Is there consensus on the restoration programme’s mission? • Have the goals and objectives been identified? • Has the restoration been planned with adequate scope and expertise? • Does the restoration management design have an annual or midcourse correction point in line with adaptive management procedures? • Are the performance indicators (the measurable biological, physical, and chemical attributes) directly and appropriately linked to the objectives? • Have adequate monitoring, surveillance, management, and maintenance programmes been developed along with the project, so that monitoring costs and operational details are anticipated, and that results of monitoring will be used to improve restoration techniques? • Has an appropriate reference system (or systems) been selected from which to extract target values of performance indicators for comparative project evaluation? • Have sufficient baseline data on the ecosystem been collected to facilitate before-and-after treatment comparisons? • To minimise the risks of failure, have critical project procedures been tested on a small experimental scale in part of the project area? • To minimise maintenance requirements, has the project been designed to make the restored ecosystem as self-sustaining as possible? • Has thought been given to how long monitoring will have to be continued before the project can be declared effective? • Have risk and uncertainty been adequately considered in project planning? During restoration • Based on the monitoring results, are the anticipated intermediate objectives being achieved? If not, are appropriate steps being taken to correct the problem(s)? • Do the objectives or performance indicators need to be modified? If so, what changes may be required in the monitoring programme? • Is the monitoring programme adequate? Post-restoration • To what extent were project goals and objectives achieved? • How similar in structure and function is the restored ecosystem to the target ecosystem? • To what extent is the restored ecosystem self-sustaining, and what are the maintenance requirements? • If all natural ecosystem functions were not restored, have critical ecosystem functions been restored? • If all natural components of the ecosystem were not restored, have critical components been restored? • How long did the project take? • What lessons have been learned from this effort? • Have those lessons been shared with interested parties to maximise the potential for technology transfer? • What was the final cost, in net present value terms, of the restoration project? • What were the ecological, economic, and social benefits realised by the project? • How cost-effective was the project? • Would another approach to restoration have produced desirable results at lower cost?

storage capacity and so on. Multiple performance measures are needed, and some of them should be robust enough and monitored often enough to detect any serious problems that may require adaptive management (Figure 32.4).

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Site selection There are three issues concerning site selection whose resolution will have major impacts on the eventual wetland after restoration: (i) the increasing

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need for larger-scale planning of projects; (ii) the issue of on-site versus off-site mitigation for the loss of wetlands; and (iii) the optimisation of site selection when there are multiple potential sites. Landscape-level planning Studies of wetland restoration in different regions of the United States (Kentula et al. 1993; Galatowitsch et al. 1998; Gwin et al. 1999) indicate that the types of wetlands being constructed often do not resemble those that existed previously or are only a subset of the types that existed previously. From a regional perspective, collective efforts to create and restore wetlands, often by a variety of different agencies and groups, are producing a landscape with a different frequency of occurrence of wetland types, even when the stated goal of these projects is to replicate wetlands that were present previously. This has resulted in the call for planning ecological restoration projects at a larger scale than the individual project, primarily at the landscape level (Baldwin et al. 1994). So far, effective means of doing this remain elusive because of economic, political and social constraints. Nevertheless, largerscale planning is a worthy and desirable goal towards which restorationists should be working (Blackwell et al. Chapter 19; Verhoeven, Chapter 12). Siting of mitigation projects In this context, mitigation involves the creation or restoration of a wetland to compensate or mitigate for wetlands lost because of a public or private development project (Marsh et al. 1996; Society of Wetland Scientists 2004). Although mitigation banking has been primarily an American concern, it is likely that it will become increasingly adopted worldwide. There are major disagreements among interested parties (developers, politicians, environmentalists, regulators and academics) on just about every aspect of mitigation (Salvesen 1994). These have ranged from ‘Is it possible to create wetlands at all?’ to ‘Who will maintain these wetlands?’ One issue that is still

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being debated is ‘Where should replacement wetlands be located?’ Two opposing answers to this question have been proposed: (i) on the same site as the wetland that will be lost; or (ii) at another site that is being developed specifically for this purpose (i.e. a ‘mitigation bank’ site). Locating restored or created wetlands for mitigation purposes on the site where the wetland was lost seems to ensure that the replacement wetland will develop under similar conditions and hence should be similar to the one lost. Unfortunately, as noted previously, there is no theoretical or practical reason to assume that this will happen automatically. A major drawback of on-site mitigation is that it results in most replacement wetlands being located in urban landscapes that do not resemble those of the lost wetlands. How this affects the functioning of these replacement wetlands and their societal value does not seem to have been studied. Another significant problem could be the long-term survival of these wetlands. An obvious disadvantage to developers of on-site, in-kind mitigation is that some of their property cannot be developed. ‘Mitigation banking’ is an alternative to on-site mitigation (Society of Wetland Scientists 2004). It involves buying areas where wetlands already exist to preserve them, or buying land on which to create or restore wetlands to replace those that will be lost. Mitigation bank sites are typically owned and managed by some kind of government agency but some are private. The main arguments against off-site mitigation are that it will result in the loss of all wetlands from developing areas, and that some kinds of wetlands for example fens or other types of wetland dependent on groundwater flows (see Money et al., Chapter 33), may be lost because it is impossible to recreate them at a mitigation bank site. The major advantages of mitigation banks are (i) that they are often of greater ecological value because the entire landscape, both uplands and wetlands, is normally restored; (ii) that the wetlands are more likely to survive in perpetuity because most will be publicly owned; and (iii) that mitigation banks are less expensive because they are not in urban areas. How far offsite a replacement wetland can be constructed

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Restoration of Wetland Environments needs careful consideration and study. Clearly, the mitigation bank used should be close enough to the wetland that will be destroyed to have similar characteristics to it. Whether on-site or off-site mitigation is the better choice can best be decided on a case-bycase basis. There does not seem to be any a priori reason to suggest that one is always preferable to the other. Two principles, however, should guide decisions about whether to allow off-site mitigation. One is that proponents of mitigation banking should be required to prove, through a pilot study if necessary, that a new wetland, comparable with the one that will be lost, can be created or restored off-site. Secondly, there should be an equal amount of expenditure for an off-site as an on-site mitigation. In other words, lower costs should not be a factor in whether to mitigate on- or off-site.

projects. Poor site selection can result in more costly projects and in a higher probability of project failure. For example, in the American upper Midwest, many more drained prairie pothole sites are available for restoration than can be restored with available resources (Galatowitsch and van der Valk 1994). In such situations, selecting sites that have the best potential for restoration becomes important. This typically means trying to identify and avoid sites that have significant impediments to restoration such as roads, railroad tracks, or other rights of way in, or adjacent to, the site, and potential problems with flooding of adjacent land. Another major consideration is the potential for being able to establish the required hydrology on a site. Many restoration sites exist in highly transformed landscapes in which the former hydrology of a site may be impossible to restore. Often, simple rule-ofthumb guidelines or nomograms (Figure 32.6) for assessing potential hydrology of sites can be developed: for example, the ratio of watershed to restored wetland area (Galatowitsch and van der Valk 1994).

Multiple potential sites There are many situations where multiple potential sites are available for creation or restoration

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Site evaluation Site evaluations should provide basic information about the site (e.g. site topography, proximity to extant wetlands, land use surrounding project site, etc.) and should identify the constraints or impediments to re-establishment of the desired wetland, including hydrological, environmental and legal problems (e.g. rights of way) and the potential for natural or spontaneous (Prach et al. 2001) re-vegetation. Potential problems with restoring or creating the wetland’s hydrology are the most common problems that will need to be evaluated. Poor soil conditions (salinity, low pH, coarse soil texture, low nutrient levels, toxic contaminants etc.) may also be important impediments to project success on some sites. Potential problems with restoring the vegetation also need to be assessed. Typically, it is a lack of suitable species on the project site and the potential for invasive species to quickly dominate the site that will be the most important problems that will need to be overcome. Environmental site evaluation Among the first steps in any planning effort should be to collect as much information about the site as feasible. An evaluation of potential problems with establishing or restoring suitable hydrology is essential, since establishment of the proper hydrology is essential for project success (Ramseier et al., Chapter 34). Where will the water come from? Where will it go? In order to develop workable planting plans, information about the topography of the basin is also essential. How deep is the water expected to be at various times of the year? Will flooding or re-flooding of the site cause any problems for property adjacent to the project? For historic projects, detailed knowledge is required about the hydrology of the wetland that previously existed on the site, or that of similar wetlands nearby. Hydrological information, unfortunately, is often limited or non-existent. Inferences about the required hydrology can often be gleaned from soils and vegetation data from

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nearby reference wetlands (Galatowitsch and van der Valk 1994). For functional projects, the hydrology simply has to be suitable for the establishment and growth of the types of wetland species needed, for example, emergent species. It does not need to be at all similar to the hydrology of natural wetlands in the same area. Restoring or creating a suitable hydrological regime is a necessary requirement, but is not itself sufficient for a successful ecological restoration project. Besides hydrology, environmental restoration may require soil amendments to improve substrate conditions, including its texture and fertility, to ensure that the substrate is suitable for the establishment and growth of wetland plants. Substrates, especially those in creation projects, are often not ideal media for plant growth. Two problems are common, low nutrient levels and little or no organic matter. Newly exposed substrates in ecological restoration projects need to be tested to see if they are adequate for plant growth (van der Valk et al. 1999). Low nutrient and organic matter levels will result in poor plant growth and high rates of plant mortality. The addition of suitable organic matter (compost, peat) can result in higher nutrient and higher soil moisture levels (van der Valk et al. 1999), which can greatly improve seed germination and seedling survival and growth rates (Figure 32.7). The

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Fig. 32.7 Mean biomass of Carex stricta grown under greenhouse conditions for three months in soils amended with 0–100% compost (van der Valk et al. 1999).

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Restoration of Wetland Environments lack of adequate attention to soil conditions can result in the total failure of vegetation establishment in an ecological restoration project. The ultimate goal of soil amendments is to ensure that an adequate supply of soil nutrients, especially phosphorous, is available to support the plant biomass and necromass that is typically found in the wetland type being established, and that soil organic matter levels are also comparable. Ideally, nutrients should be added in the form of organic matter amendments that release the required nutrient as the organic matter decomposes. The possible presence of toxic compounds, especially long-lasting pesticides and heavy metals, also needs to be considered during a site evaluation (Lyle 1987; Harris et al. 1996; Marburger et al. 1999; Ramseier et al., Chapter 34). This is especially true for created wetlands and restored wetlands on sites that were formerly farmed or used for industrial purposes or mining. For example, more than 20 years after they were banned, residual organochlorine pesticides (Table 32.3) were found in the soils and fish of some units of the restored Emeralda Marsh, in central Florida, that had previously been farm land (Marburger et al. 1999). In 1998, nearly a thousand white pelicans died, apparently after eating pesticide-contaminated fish in a newly restored wetland, the North Shore Restoration Area of Lake Apopka in central Florida. This site had been farmed for many years, and it is believed that, as with the nearby Emeralda Marsh

Table 32.3 Organochlorine pesticide contaminants (µg kg−1) of Everglades muck sediments in 1996 in various units of the Emeralda Marsh Conservation Area, Florida, USA. Application of these pesticides was banned in 1973. (Data from Marburger et al. 1999.) Marsh unit

n

DDD DDE

Eustis 18 3900 Knight North 10 ND* Knight South 3 ND Long 36 1300 * ND = none detected.

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2800 ND ND 800

DDT Dieldrin Toxaphene 570 ND ND 60

390 ND ND 200

39 400 1 600 1 500 12 800

745

complex, significant levels of organochlorine pesticides were still present when the site was re-flooded. The pesticides quickly found their way into the developing food chains. When water levels in the wetland were drawn down in late 1998; fish were concentrated in pools, and this attracted pelicans from around the region, with disastrous consequences. Biological site evaluation What impediments are there to the establishment of the desired vegetation? What is the potential for natural re-vegetation? Are the seeds of wetland species present in the relict seed bank? Are wetland species growing on or adjacent to the project? Whether there is any need to introduce species will depend primarily on whether plant propagules of the desired species are still present or can still readily reach the project site. Riverine, lacustrine and some coastal sites may still receive significant inputs of seeds and other propagules from adjacent wetlands (Gordon and van der Valk 2003) and waterfowl may carry seeds to isolated wetlands (Mueller and van der Valk 2002). This seed rain, however, can also contain seeds of unwanted species and thus can be both a blessing and curse for historic creation or restoration projects. For example, Mitsch et al. (1998) reported that many species were able to colonise two wetlands created along the Olentangy River in Ohio because their seeds were carried into the wetland by floodwater or were pumped in. On the other hand, Galatowitsch and van der Valk (1995, 1996) found that many guilds of wetland species were not readily re-colonising restored prairie pothole wetlands (Figure 32.8) because of their isolation from extant wetlands. For historic projects, rapidly establishing suitable species is essential because otherwise the site may become quickly dominated by weed species. For restoration projects, it is sometimes assumed that seeds of wetland species are still present in the soil and that adding water is all that is needed to restore the wetland. Studies of the seed banks of drained wetlands suggest that relict seed banks may not

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Wet prairie species

Deep emergent species EAGLE LAKE GROVERS EAST IA148 MN176 IA149 MN276 IA027 IA140 IA150 IA125 S0008 IA092 IA016 IA017 IA172 S0007 IA147 S0017 IA097 MN054 MN268 S0002 IA052 IA160

DOOLITTLE JANSSEN EAST IA148 MN176 IA149 MN276 IA027 IA140 IA150 IA125 S0008 IA092 IA016 IA017 IA172 S0007 IA147 S0017 IA097 MN054 MN268 S0002 IA052 IA160 0

3

6

9

12

15

18

21

0

24

Sites

Sedge meadow species

1

2

3

4

5

6

10

12

Submersed aquatic species

CAYLER BIEBER IA148 MN176 IA149 MN276 IA027 IA140 IA150 IA125 S0008 IA092 IA016 IA017 IA172 S0007 IA147 S0017 IA097 MN054 MN268 S0002 IA052 IA160

GOOSE LAKE VENTURA IA148 MN176 IA149 MN276 IA027 IA140 IA150 IA125 S0008 IA092 IA016 IA017 IA172 S0007 IA147 S0017 IA097 MN054 MN268 S0002 IA052 IA160 0

6

12

18

24

30

36

42

48

0

2

4

6

8

Number of species observed Shallow emergent species GOOSE LAKE JANSSEN WEST IA148 MN176 IA149 MN276 IA027 IA140 IA150 IA125 S0008 IA092 IA016 IA017 IA172 S0007 IA147 S0017 IA097 MN054 MN268 S0002 IA052 IA160

High quality

Typical

0

3

6

9

12

15

18

21

24

Number of species observed Fig. 32.8 The number of species in various guilds of wetland species in 22 restored prairie pothole wetlands. The number of species in high and typical quality natural prairie potholes is shown in the first two bars of each panel. (From Galatowitsch 1993.)

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Restoration of Wetland Environments play a significant role in the re-vegetation of restored wetlands (Weinhold and van der Valk 1989; Brown 1998). If the project will require planting selected wetland species, three issues typically arise. These are: (i) the source of plant materials; (ii) weed control measures; and (iii) planting techniques. The control of grazing animals, especially geese, can also be an important issue at some sites (see also Van Den Wyngaert and Bobbink, Chapter 14; Ross and Murkin, Chapter 35). Commercial nurseries in some regions can supply seeds, seedlings and adult plants of at least selected wetland species. However, the quality of plant material that is supplied by commercial nurseries can vary widely. For example, seeds are sometimes collected from local wetlands by native plant nurseries. The viability of this seed, however, is rarely tested and can be very low (van der Valk et al. 1999). The exact form of plant material needs to be specified when bids are solicited from nurseries (Garbisch 1990): for example the age or size of the plants, the minimum number of stems or shoots per pot, size of pot in which they are to be grown and type of soil in which they are to be grown. The nursery must be able to certify that these conditions were met when the plants were grown. They may also be required to prove that the seedlings are disease free. Because its use may result in the loss of local genotypes and the introduction of new genotypes, controversy sometimes surrounds the use of commercial sources of plant material when it is not of local provenance. Restorationists have often overlooked the possibility of genetic contamination of native populations by gene flow from introduced genotypes. Lessica and Allendorf (1999) provide an overview of population genetics, including local adaptations, genetic drift, gene flow and hybridisation, as it relates to ecological restoration projects. Handel et al. (1994), Jones (2003) and McKay et al. (2005) also provide good overviews of this important topic as do earlier papers by Berger (1991, 1993). Conventional wisdom suggests that it is best to use local plant material if it is available

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(Handel et al. 1994). What local means in this context has not always been well defined (McKay et al. 2005). It varies from species to species because of differences in potential dispersal distances of pollen and seed. Local genotypes should be used for projects where environmental conditions and disturbance regimes are similar to those in extant wetlands, especially for historic projects. For projects where environmental conditions, especially soils, or disturbance regimes are very different from those in extant wetlands, local genotypes may not be a good choice (Lessica and Allendorf 1999). For example, Bremholm (1993) found that it was impossible to establish Carex spp. in a created wetland because of the site’s coarse soils. On highly disturbed sites, plant varieties are needed that can become established and survive under extreme conditions. Although it may be possible to find suitable local genotypes growing in nearby disturbed wetlands (Handel et al. 1994), commercial cultivars are often the only plant material that can be established on such sites. Mixtures of local genotypes and cultivars may also be useful on disturbed sites. There are, however, potential genetic consequences of planting such mixtures that should be carefully evaluated. Lessica and Allendorf (1999) provide the following guidelines for selection of plant materials: • Avoid strongly selected cultivars. • Initiate new populations with as many genotypes and as much genetic variation as possible, especially when disturbances have been severe. • Use local genotypes when feasible, especially for large sites. • Use phenotypically plastic species with wide ecological amplitudes when feasible. • Use selfing species or those with low dispersal capabilities to minimise the genetic contamination of resident populations. Another potential way to establish wetland plant species is to use relict (Wetzel et al. 2001) or donor seed banks (van der Valk and Pederson 1989; Brown and Bedford 1997). Donor seed banks use surface soil taken from an existing wetland and spread it on the substrate of the created or restored wetland. This soil contains seeds and

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other propagules of plant species as well as a host of microorganisms and invertebrates. A major problem with this technique is locating an adequate source of donor soil. This may be available only at mitigation sites where a native wetland is to be destroyed. Even then, the soil of the donor wetland may be unsuitable if it is contaminated with the seeds of unwanted species. An additional problem is that the seed of some desired species may not be present in the seed bank, for example short-lived seeds of many wetland forest species (van der Valk and Pederson 1989). Nevertheless, Brown and Bedford (1997) concluded that it can be an effective technique for establishing wetland species and that it can reduce the recruitment of undesirable species. During the environmental and biological restoration stages (Figure 32.4) when the project is actually being implemented, numerous ecological decisions will still need to be made, for example, about the timing of plant establishment. The correct timing of plant introductions is determined by seed germination requirements (temperature, moisture, salinity) or by the water depth or soil moisture tolerances of their seedlings. Differences in soil moisture, water depth and soil temperatures can significantly affect species recruitment (Seabloom et al. 1998). Most ways of establishing plants in wetland projects have been adapted from forest and grassland planting techniques. These are described in various restoration manuals such as Packard and Mutel (1997) and Harris et al. (1996). The success of these efforts to establish plant species will often be determined by how well the hydrology of the new wetland is understood and controlled. Failure to establish plants often results from too rapidly rising or falling water levels, which thereby either drown seeds or newly established plants, or kill plants by desiccation. An ability to control soil moisture and water levels during the vegetation restoration phase will significantly improve the chances of successful plant establishment. One of the major problems in many historic ecological restoration projects is the establishment of unwanted weedy species (D’Antonio and Meyerson 2002; Zedler and Kercher 2004;

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Ramseier et al., Chapter 34). Two techniques can be used to reduce this problem. If the species are being established using seed, a matrix species can be included in the seed mix. This is typically an annual grass that cannot survive flooding. The seed mix is sown while the wetland substrate is kept moist but not flooded, which allows the matrix as well as most wetland species to germinate and grow. The matrix species, because it grows quickly and initially dominates the vegetation, hinders the colonisation of unwanted species and gives the desirable species, sown at the same time, a better chance to become established. After the wetland species have become established, the new wetland is re-flooded, thereby eliminating the matrix species and allowing the clonal spread of wetland species and the germination of submerged aquatics. The other technique is the eradication of the seedlings of unwanted species. This can be done by hand, mechanically or chemically (see Pieterse and Murphy (1990) for a detailed discussion of aquatic weed control techniques). Seedlings in a newly restored or created wetland often attract grazing animals to the site, especially geese. Within hours these grazers can destroy the vegetation that the restorationist has spent days or weeks trying to establish. Again, there should be some contingencies for dealing with grazing animals, especially in the early stages of vegetation restoration. Restoration plan The data from the site evaluation on basin topography, hydrology, environmental conditions and potential for natural re-vegetation are used to develop a restoration plan. This plan describes what will need to be done to establish or restore the wetland on the site. It should outline all of the problems identified and the proposed ways to overcome them. Restoration plans can vary from a page or two for small, routine projects to multivolume tomes running to thousands of pages. For large projects, they need to be reviewed by appropriate professionals to ensure that there have been no oversights or mistakes. The plan needs to be detailed enough, including cost estimates, so that

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Restoration of Wetland Environments it can be used to prepare bid documents for all the work that will need to be contracted out. Monitoring and adaptive management Because of our imperfect knowledge about how to create or restore wetlands and to make adaptive management feasible, it is necessary to monitor ecological restoration projects during the environmental and biological restoration stages (Kentula 2000). Monitoring is also needed to improve future projects by providing data on what worked and what went wrong. The partial failure of ecological restoration projects is common (Kusler and Kentula 1990). There are many reasons for these failures, including inadequate scientific knowledge, poor project design, inadequate supervision during the implementation, inadequate site preparation, invasion of weed species and extirpation of newly planted vegetation. Many problems can be corrected easily, if they are discovered early enough. Adaptive management allows for the iterative refinement of management plans as more information about the developing wetland is acquired (Ringold et al. 1996; Weinstein et al. 1997; Schreiber et al. 2004). Common management manipulations include adjusting the hydrology, replanting, re-grading part of the site and improved weed control measures. The level of monitoring required depends on the nature of the project. It can range from an occasional trip to the site to see what has happened, to longterm, quantitative studies (Goldsmith 1991; Spellerberg 1991). The latter are often mandated by government agencies if the project is mitigating for the loss of a wetland. At a minimum, some photographs of the site should be taken regularly and a visual examination of dykes, control structures and the condition the vegetation should be made periodically (Galatowitsch and van der Valk 1994). In spite of its importance to the ultimate success of a project, monitoring is not common (Kusler and Kentula 1990; Grayson et al. 1999). Claims made for the success of projects where there has been no monitoring are usually inflated and are typically based solely on acreage

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or number of wetlands. The compositional and functional attributes of these unmonitored wetlands, however, may bear little resemblance to those of natural wetlands. Two excuses are typically given for not doing any monitoring: (i) it is not feasible because of a lack of personnel or funds; (ii) the money needed to do monitoring is better spent on additional projects. Claims about the success of wetland projects where little or no monitoring was done should be treated with a great deal of scepticism (Kusler and Kentula 1990; Galatowitsch and van der Valk 1994). Without monitoring, restoration ecology will not develop as a scientific discipline, and disillusionment and scepticism will grow about our ability to design and implement successful projects. To aid in designing effective monitoring plans, the following guiding principles should be considered: • Before the project starts, monitoring specifics need to be agreed on by all parties. • Proposed monitoring schemes need to be realistic, reliable and affordable. • Performance measures to be monitored need to be strongly linked to project goals. • The monitoring plan needs to take into account site spatial and temporal variability. • Multiple performance measures of site condition should be monitored. • Some performance measures should be reliable enough to be used for adaptive management decision making. Project assessment Project assessments are commonly done for three reasons: (i) they are required to demonstrate that the project has met any legal requirements imposed on it; (ii) so that the contractor(s) who carried out the project can get paid; and (iii) to document what was learned about the restoration of wetlands by this particular project. The last of these is ultimately the most important. Assessments are made by comparing the performance measures developed at the start of the project with relevant data collected during the course of the project. How exactly such project assessments are done is poorly documented and much

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more research needs to be done on this important stage of restoration management. Some questions that need to be answered during the design of an assessment are: • How will you decide the project was or was not a success? How many performance measures have to be met? • Which are the most important performance measures? Should performance measures be combined into a single index of success? • Who will do the project assessment? When will it be done? • How will uncontrollable events like poor weather conditions or construction delays be handled? Monitoring data and evaluation results need to be incorporated into a formal project report that is available to interested parties. At a minimum, this means putting the monitoring data and assessment report on the web. It is only by sharing results from restoration and creation projects that it will be possible to improve the probability of success of future projects.

ACK N O W L E DGE M E N T S I would like to thank Elizabeth Gordon for the opportunity to co-teach restoration ecology at the Central University of Caracas. Discussions with her and the students in this course about both the theoretical and practical aspects of restoration greatly influenced the contents of this chapter. I would also like to acknowledge the support of several funding agencies which have supported our various studies of wetland restoration, including the US Environmental Protection Agency, the South Florida Water Management District and the Delta Waterfowl and Wetlands Research Station.

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Berger J.J. 1991. A generic framework for evaluating complex restoration and conservation projects. Environmental Professional 13, 254–262. Berger J.J. 1993. Ecological restoration and nonindigenous species: a review. Restoration Ecology 1, 74–82. Bremholm T.L. 1993. Evaluation of Techniques for Establishing Sedge Meadow Vegetation. MS thesis, Iowa State University, Ames, IA. Broome S.W. 1990. Creation and restoration of tidal wetlands of the southeastern United States. In: Kusler J.A. and Kentula M.E. (editors), Wetland Creation and Restoration: The Status of the Science. Island Press, Washington, DC, pp. 37–72. Brown S.C. 1998. Remnant seed banks and vegetation as predictors of restored marsh vegetation. Canadian Journal of Botany 76, 620–629. Brown S.C. 1999. Vegetation similarity and avifaunal food value of restored and natural marshes in northern New York. Restoration Ecology 7, 56–68. Brown S.C. and Bedford B.L. 1997. Restoration of wetland vegetation with transplanted wetland soil: an experimental study. Wetlands 17, 424–437. Christensen N.L., Bartuska A.M., Brown J.H., Carpenter S., D’Antonio C., Francis R., Franklin J.F., MacMahon J.A., Noss R.F., Parsons D.J., et al. 1996. The report of the Ecological Society of America Committee on the scientific basis for ecosystem management. Ecological Applications 6, 665–691. Clements F.E. 1916. Plant Succession: An Analysis of the Development of Vegetation. Publication 242, Carnegie Institution, Washington. Clements F.E. 1928. Plant Succession and Indicators. Wilson, New York. Cole A.C. 1999. Ecological theory and its role in the rehabilitation of wetlands. In: Streever W. (editor), An International Perspective on Wetland Rehabilitation. Kluwer, Dordrecht, pp. 265–275. Committee on Mitigating Wetland Losses 2001. Compensating for Wetland Losses Under the Clean Water Act. National Academy Press, Washington, DC. Committee on the Restoration of Aquatic Ecosystems 1992. Planning and evaluating aquatic ecosystem restoration. In: Restoration of Aquatic Ecosystems. National Research Council, Washington, DC. Craft C., Megonical P., Broome S., Stevenson J., Freese R., Cornell J., Zheng L., and Sacco J. 2003. The pace of ecosystem development of constructed Spartina alterniflora marshes. Ecological Applications 13, 1417–1432. Dahm C.N. 1995. Special issue: Kissimmee River restoration. Restoration Ecology 3, 145–238.

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Shaffer P.W., Kentula M.E. and Gwin S.E. 1999. Characterization of wetland hydrology using hydrogeomorphic classification. Wetlands 19, 490–504. Simenstad C.A. and Thom R.M. 1996. Functional equivalency trajectories of the restored Gog-LeHi-Te estuarine wetlands. Ecological Applications 6, 38–56. Society of Wetland Scientists 2004. Wetland Mitigation Banking Viewed 14 June 2006 at http://www.sws.org/ wetlandconcerns/banking.html Spellerberg I.F. 1991. Monitoring Ecological Change. Cambridge University Press, Cambridge, 334 pp. Sprugel D.G. 1991. Disturbance, equilibrium, and environmental variability: what is ‘natural’ vegetation in a changing environment? Biological Conservation 58, 1–18. Streever W. (editor) 1999. An International Perspective on Wetland Rehabilitation. Kluwer, Dordrecht. Toth L.A. 1995. Principles and guidelines for restoration of river/floodplain ecosystems – Kissimmee River, Florida. In: Cairns J. (editor), Rehabilitating Damaged Ecosystems, (2nd edition). Lewis Publishers, Boca Raton, FL, pp. 49–73. Toth L.A. and Anderson D.H. 1998. Developing expectations for ecosystem restoration. Transactions of the North American Wildlife and Natural Resources Conference 63, 122–134. Urban D.L. and Shugart H.H. 1992. Individual-based models of forest succession. In: Glenn-Lewin D.C., Peet R.K. and Veblen T.T. (editors), Plant Succession: Theory and Prediction. Chapman and Hall, London, pp. 249–292. van Andel J. and Aronson J. 2006. Restoration Ecology. Blackwell, Oxford, UK. van Andel J., Bakker J.B. and Grootjans A.P. 1993. Mechanisms of vegetation succession: a review of concepts and perspectives. Acta Botanica Neerlandica 42, 413–433. van der Valk AG. 1981. Succession in Wetlands: a Gleasonian approach. Ecology 62, 688–696. van der Valk A.G. 1982. Succession in temperate North American wetlands. In: Gopal B., Turner R.E., Wetzel R.G. and Whigham D.F. (editors), Wetland Ecology and Management. International Scientific Publications, Jaipur, India, pp. 169–179. van der Valk A.G. 1985. Vegetation dynamics of prairie glacial marshes. In: White J. (editor), Population Structure of Vegetation. Junk, The Hague, pp. 293–312. van der Valk A.G. 1987. Vegetation dynamics of freshwater wetlands: a selective review of the literature.

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33 Replumbing Wetlands – Managing Water for the Restoration of Bogs and Fens R U S S P . MONE Y 1 , B R Y AN D . WH EELER 2 , AND Y J . BAI RD 3 AND A . L O U I SE H EATH WAI TE 4 1Natural England, Reading, UK of Animal and Plant Sciences, University of Sheffield, Sheffield, UK 3School of Geography, University of Leeds, Leeds, UK 4Centre for Sustainable Water Management, Lancaster Environment Centre, University of Lancaster, Lancaster, UK

2Department

IN T R O D U CT ION Restoration of wetlands Population pressure, and the requirement for ‘dry’ land which can be settled and cultivated, has meant the drainage and loss of large areas of wetland, particularly across north-west Europe (Verhoeven 1992; Lappalainen 1996). Consequently, residual wetland is often highly valued by conservationists, though many sites remain vulnerable to damage, either directly or indirectly, from ongoing activities such as drainage, groundwater abstraction, forestry and peat extraction (Williams 1990; Bragg et al. 1992; Fojt 1995). Wetland restoration can entail the repair of damaged but extant wetland but also the re-creation of wetland habitat in areas where essentially it has been lost completely (van der Valk, Chapter 32). For much of the twentieth century, restoration initiatives, where they have occurred, have been driven principally by the nature conservation objectives of governments and non-government organisations within

The Wetlands Handbook Edited by Edward Maltby and Tom Barker © 2009 Blackwell Publishing Ltd. ISBN: 978-0-632-05255-4

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individual countries. More recently, however, wetland restoration has received additional impetus from international initiatives such as the European Union Habitats and Species Directive, the European Wild Birds Directive and the Convention on Biological Diversity (HMSO 1994; Fojt 1995; Raeymaekers 1997). Whilst nature conservation has been the primary driving force behind much wetland restoration, increasingly it is being seen as just one of a number of valuable ‘functions’ that wetlands can perform for human society (Maltby 1998). For example, there is growing appreciation of the role wetlands can play in regulating water supply (Burt 1995), influencing water quality (Ross 1995), and managing river flooding and coastal erosion (Williams 1990; Joyce and Wade 1998). Certain types of wetland can support livestock (Benstead et al. 1997) whilst others provide a renewable source of harvestable products (Andrews 1992). There is even growing interest in the potential for more productive wetland to be managed for biofuel production. Globally, peat-forming wetlands are a significant sink for atmospheric CO2, which may have important implications for global carbon cycling and climate change (Martikainen 1996; Tuittila et al. 1999). Consequently, amidst current concerns about global warming, flooding, coastal erosion, unsustainable agricultural practices and biodiversity loss, there is an increasing realisation

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that current land-use practice in relation to wetlands is both suboptimal and unsustainable. After a long history of drainage and reclamation, a wetland renaissance is under way, and restoration undoubtedly has a key role to play. Moreover, the restoration of wetland will usually require some degree of ‘replumbing’, from repair of minor ‘leaks’ to reinstatement, or replacement, of entire water supply mechanisms. This chapter is concerned with the restoration of ‘telmatic wetlands’, that is, wet, but essentially terrestrial systems (or ‘mires’ (Gore 1983a)), as opposed to ‘aquatic-wetlands’ which might loosely be described as shallow water systems largely without emergent plant cover (Wheeler 1995, 1999a). In many instances, telmatic and aquatic wetlands occur in juxtaposition as part of a wetland complex, and whilst the distinction may seem arbitrary, it is nevertheless quite important because their ecology and restoration requirements are different. The ‘wetness’ of wetlands The common ingredient of all natural wetlands is an ‘excess’ of water, or at least a hydrological balance adequate to create conditions in which the surface is usually waterlogged for at least part of the year. The source of the water, its quantity and quality and the mechanism by which it is delivered to the wetland combine to influence wetland development and character, giving rise to the wide spectrum of different wetland types that occur in the landscape (Gore 1983b; Moore 1984; Moore and Bellamy 1974; Wheeler 1995). Wet conditions may result from impeded drainage, high rates of water supply or both. Water supply may consist of telluric water (i.e. water that has had some contact with mineral ground such as riverwater, surface runoff or groundwater discharge) or meteoric water (i.e. precipitation). It has long been recognised that differences in topographic situation and water supply mechanism profoundly influence wetland type. Such considerations formed the basis of the early but long-standing systems of wetland classification. Von Post and Granlund (1926) subdivided

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wetlands into three types: ombrogenous, developed under the exclusive influence of precipitation; topogenous, irrigated by telluric water that naturally collects on flat ground and topographic hollows; and soligenous, developed on slopes and kept wet by a supply of telluric water. On the basis of floristic variation in Swedish wetlands, Du Rietz (1949, 1954) emphasised that mires could be divided into areas fed almost exclusively by precipitation (ombrotrophic or rain-fed) and those in which water supply was supplemented by telluric water (minerotrophic or rock-fed). These early concepts are important because they broadly correspond with major habitat differences still recognised today. Wetland scientists have come to use the word fen as a synonym for minerotrophic wetlands and bog to refer to ombrotrophic examples (these concepts have been reviewed by Wheeler and Proctor (2000)). Restoration objectives Given the wide range of wetland types that exist, it follows that restoration objectives can be equally diverse. The precise objectives for individual sites are often determined by the perceived conservation desirability of a particular wetland habitat or species and do not necessarily correspond to the ‘natural’ condition of the wetland. In many cases, objectives are linked with some recorded former state of the wetland prior to damage or dereliction of traditional vegetation management practices (Wheeler 1996). In other instances, where most vestiges of the original wetland have been destroyed, conservationists can have a ‘clean slate’ upon which to work. For example, in the UK large areas of ‘reed bed’ have been created from agricultural land and peat workings with the specific objective of generating habitat for bittern (Botaurus stellaris), one of the UKs rarest wetland birds (Hawke and José 1996). Because of this range of objectives, practical wetland restoration initiatives can take many forms, from the blocking of drains in sites damaged in a relatively minor way to the wholesale conversion of large tracts of farmland to reed bed. An insight into the broad scope of wetland

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Replumbing Wetlands restoration activity is provided by Wheeler et al. (1995), Wheeler and Shaw (1995a), Malterer et al. (1998), Pfadenhauer and Klotzli (1996), Hawke and José (1986), RSPB et al. (1997) and (Ramseier et al., Chapter 34). Restoration projects aimed at achieving a particular endpoint will stand a greater chance of success if the environmental requirements of the target habitats or species are known. For the vegetation, the critical zone of influence lies at or near the surface of the wetland. Consequently, the challenge for wetland restorers is to be able to identify the required surface conditions and to restore, or generate, a hydrological mechanism capable of delivering them (van Wirdum 1995). The aim of this chapter is to explore the relationship between aspects of the ecology and hydrology of wetlands and the significance of this to restoration planning and practice. Consideration is given to key features of the surface environment that influence wetland vegetation composition and the way in which these features are linked to wetland hydrodynamics, which can be manipulated as part of the wetland restoration process.

G E NE RAL CO N S ID E R AT IO N S O F T HE RE P L U M B IN G OF W E T L AN DS

rare habitats (e.g. Cirsium dissectum – Molinia caerulea fen meadow, in western Europe). The loss of ‘favourable’ hydrological conditions at the soil surface can be a product of one or more of three main processes: reduction of water supply, increased water loss or disturbance of hydroregulatory processes internal to the wetland such as acrotelm function (Heathwaite et al. 1993). Before any attempt at replumbing wetlands can be made, the cause of dryness needs to be determined. Traditionally, hydrologists have estimated water budgets for individual sites, but it is important to recognise that the ecohydrological value of these depends considerably upon their scale and focus and some water budget studies appear to reach misleading conclusions. For example, a water budget for a spring-fed wetland in Norfolk, England, calculated by Adams et al. (1994), indicated that the proportionate contribution of surface water was greater than that of groundwater. However, the catchment-based estimate of the contribution of ‘surface water’ appears mainly to represent just water that passes through the wetland in a stream which probably has little, if any, direct influence on the seepage slopes. If water budget approaches are to be used, it is important that they can be related to the key features of ecological interest. Water sources

The need for replumbing The recognition of a need for replumbing of wetlands implies some deterioration from a former desirable state. In most cases this is manifest in drier conditions at the soil surface than was once the case, though in some cases restoration of flooded wetlands to a less flooded state may also be seen as desirable. Although emphasis is often – and legitimately – placed upon the problems of reduced water tables, it is important to recognise that permanently high, above-surface, water levels are often associated with rather species-poor wetland habitats. Some rich examples of wetland vegetation have apparently been produced by the partial drainage of once wetter sites, including some types that have international protection as

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One of the main causes of drying of some wetland sites is a reduction of rates of water supply, from one or more sources. For example, lowering of water tables in aquifers, as a consequence of borehole pumping, can reduce groundwater inputs into seepage-fed sites (Harding 1993), whilst in floodplain sites physical separation of wetlands from rivers (by embankments or drainage of intervening land) can prevent, or greatly restrict, ingress of riverwater. One solution to this problem is to try to reinstate former water supply mechanisms (when these are known), but sometimes this is impractical and alternative or supplementary water sources must be considered. This requires an understanding of the eco-hydrological role of specific water sources, and depends also

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upon the specific goals of replumbing initiatives (See Section 2, especially Grootjans et al., Chapter 8). The importance of individual water sources to wetland sites – and hence the possibility of some substitution for them by replumbing – depends critically upon their eco-hydrological function within the wetland, rather than just on their quantitative contribution. This again highlights a potential limitation of simple water balance studies in some sites. From an eco-hydrological perspective, a distinction must be made between water sources that are important to the character and ‘normal’ water table of a wetland and those that are largely incidental to it. For example, in the case of certain soligenous wetlands, groundwater springs may be the primary source of water and its quantity and quality

determine surface conditions, (e.g. spring flow elevates oxidation-reduction potentials, whilst co-precipitation of P onto calcite may also occur). The specific source in this case is usually irreplaceable. In other instances, the original source potentially could be replaced or supplemented by water of similar chemical characteristics from a different source. For example, in a floodplain context, plant communities similar to those of seepage slopes can develop on floating mats, irrigated by a lateral penetration of base-rich riverwater (Figure 33.1; van Wirdum 1991). The approach of utilising alternative water sources may be particularly viable where the character of the substratum (e.g. brackish clay) strongly determines chemical conditions. In other instances, potential water sources may actually provide no direct supply to the wetland. This could occur, for example,

Fig. 33.1 Hydrological dynamics of a floodplain fen, with a solid (top) and semi-floating (bottom) surface.

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Replumbing Wetlands where groundwater in an underlying aquifer is confined by a low permeability material such as clay or even poorly conductive peat. It may also occur despite hydrological continuity between the aquifer and the wetland. However, in this case, despite having no direct supply role, aquifer water levels may have an important supporting function, and a reduction in water level (e.g. due to abstraction) could result in enhanced vertical water loss from the wetland (Schouwenaars and Vink 1992). Water losses Many wetland sites have become damaged by increased rates of water loss. The most widespread cause of this – direct drainage – may appear to be the easiest to remedy, by blocking drainage ditches. Drainage accelerates the rate of decomposition of the peat which in turn produces changes in the hydrology, morphometry and ecology of a wetland. Such changes may not be reversible (Heathwaite et al. 1993) – even with ditch blocking. The intentional outcome of wetland drainage is lowering of the water table. The efficacy of drainage is related to the depth of ditching, distance between ditches and the hydraulic conductivity of the peat (Boelter 1972). The drawdown is greatest near the ditch, and diminishes with distance. Boelter (1972) found ditches to be effective for distances of up to 50 m in fibric peat in Minnesota, but ineffective beyond 5 m in more decomposed peat where the hydraulic conductivity is lower. Similar restrictions on drain impact exist for other peat types. For wetlands drained for afforestation, Pyatt (1987) and Robinson (1985) showed that the drying of the wetland surface and disruption of the water table and surface flow patterns were irreversible. Where field drains have been installed within the wetland, for example, in large areas of the Somerset Levels in England (Heathwaite 1990), simple ditch blocking may be only partially successful. Even when sites have not been drained directly, they can be much affected by drainage and other activities within their catchment, especially

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when this results in a reduction of groundwater tables in an aquifer that underlies, and is in hydraulic continuity with, the wetland site. A particular case of diffuse drainage damage can occur in some wetlands where the water table within the substratum is perched above the level of the regional water table. This occurs particularly – and naturally – in ombrogenous peatlands, in which the ombrotrophic peat has accumulated in association with a perched water mound derived from precipitation, but it also occurs – though normally unnaturally – in some fen systems where a residual block of fen peat is perched above its surroundings. In both situations, the height of the perched water table is partly a function of the basal area of the wetland deposit, and reduction of the basal area (e.g. by peat extraction or agricultural conversion of part of the wetland), may lead to increased drainage and a lowering of the perched water table. This is often just expressed as a drier drawdown zone around the periphery of the upstanding deposit, but in some cases – especially when combined with some surface drainage – it can lead to a reduction of water table across the whole area of the wetland (Bragg 1995; Wheeler 1996). Hydrological self-regulation Some wetlands appear to have an important capacity for internal regulation of their hydrodynamics. This has generally received little detailed investigation, though it may be critical to maintaining the character of certain types of wetlands. In some cases, these processes may also help mitigate the impact of external controls upon the water regime. Acrotelm function Some peatlands can be described as two-layered or diplotelmic systems with an uppermost acrotelm and an underlying catotelm layer (Ingram and Bragg 1984). The acrotelm is perhaps of greatest importance in ombrogenous mires, where the term is most frequently used to describe the layer of living plants (principally Sphagnum species)

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and proto-peat, which forms a relatively thin skin (typically