Climate Change Impacts on Soil Processes and

Developments in Soil Science

Climate Change Impacts on Soil Processes and Ecosystem Properties

Developments in Soil Science Series editor William R. Horwath

Developments in Soil Science

Climate Change Impacts on Soil Processes and Ecosystem Properties Edited by William R. Horwath

Department of Land, Air and Water Resources University of California, Davis, CA, United States

Yakov Kuzyakov

Department of Soil Science of Temperate Ecosystems Department of Agricultural Soil Science University of Go¨ttingen, Go¨ttingen, Germany

Elsevier Radarweg 29, PO Box 211, 1000 AE Amsterdam, Netherlands The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States Copyright © 2018 Elsevier B.V. All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Details on how to seek permission, further information about the Publisher’s permissions policies and our arrangements with organizations such as the Copyright Clearance Center and the Copyright Licensing Agency, can be found at our website: www.elsevier.com/permissions. This book and the individual contributions contained in it are protected under copyright by the Publisher (other than as may be noted herein). Notices Knowledge and best practice in this field are constantly changing. As new research and experience broaden our understanding, changes in research methods, professional practices, or medical treatment may become necessary. Practitioners and researchers must always rely on their own experience and knowledge in evaluating and using any information, methods, compounds, or experiments described herein. In using such information or methods they should be mindful of their own safety and the safety of others, including parties for whom they have a professional responsibility. To the fullest extent of the law, neither the Publisher nor the authors, contributors, or editors, assume any liability for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions, or ideas contained in the material herein. Library of Congress Cataloging-in-Publication Data A catalog record for this book is available from the Library of Congress British Library Cataloguing-in-Publication Data A catalogue record for this book is available from the British Library ISBN: 978-0-444-63950-9 For information on all Elsevier publications visit our website at https://www.elsevier.com//books-and-journals

Publisher: Candice Janco Acquisition Editor: Candice Janco Editorial Project Manager: Emily Thomson Production Project Manager: Prem Kumar Kaliamoorthi Designer: Greg Harris Typeset by TNQ Technologies

Contents List of Contributors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ix 1

Soils, Climate, and Ancient Civilizations . . . . . . . . . . . . . . . . . . . . 1 Eric C. Brevik, Jeffrey A. Homburg, Jonathan A. Sandor Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 The Use of Soils in Archaeology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3 Studies at Ancient Sites to Understand Soils . . . . . . . . . . . . . . . . . . . . . . 4 Soil Knowledge and Management in Early Civilizations . . . . . . . . . . . . . 7 Effects of Ancient Agriculture on Soils and Societies . . . . . . . . . . . . . . 12 Climate Change and Ancient Cultures . . . . . . . . . . . . . . . . . . . . . . . . . 17 Concluding Statements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20 Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21

2

SoilePlanteAtmosphere Interactions: Ecological and Biogeographical Considerations for Climate-Change Research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29 Lucas C.R. Silva, Hans Lambers Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29 Terrestrial Life as a Stabilizing Climatic Force . . . . . . . . . . . . . . . . . . . 31 Contemporary Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34 Simplifying Complexity at the SoilePlanteAtmosphere Interface . . . . . 39 Gaps in Knowledge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42 Conservation and Management Opportunities . . . . . . . . . . . . . . . . . . . . 45 Final Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48 Further Reading. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

3

The Potential for Soils to Mitigate Climate Change Through Carbon Sequestration. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61 William R. Horwath, Yakov Kuzyakov Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61 Humanities Reliance and Impact on Soils . . . . . . . . . . . . . . . . . . . . . . . 61 Soil Organic Carbon Balance and Management to Sequester Carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63 Animal Manures Sequester Soil Organic Carbon . . . . . . . . . . . . . . . . . . 65 Potential to Sequester Soil Organic Carbon. . . . . . . . . . . . . . . . . . . . . . 68 Soil Organic Carbon Sequestration to Address Climate Change . . . . . . . 78 Sequestering Soil Organic Carbon Requires N . . . . . . . . . . . . . . . . . . . 79

v

vi

Contents Atmospheric Composition and Climate Change Impacts on Soil C Sequestration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Research Needs in Soil Organic Carbon Sequestration. . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Further Reading. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . .

. . . .

. . 81 . . 82 . . 83 . . 92

4

Role of Mineralogy and Climate in the Soil Carbon Cycle . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93 Katherine Heckman, Craig Rasmussen Mineralogy, Weathering, and the Inorganic C Cycle . . . . . . . . . . . . . . . 93 Climate, Mineral Assemblage, and Soil Organic Carbon Are Intrinsically Linked Through Weathering Processes . . . . . . . . . . . . . 96 Mineral Stabilization of Soil Organic CdBonding Mechanisms . . . . . 102 Mineral Stabilization of Soil Organic CdField and Lab-Based Evidence . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 104 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 107 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 107

5

Impacts of Climate Change on Soil Microbial Communities and Their Functioning . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111 Franciska T. de Vries, Robert I. Griffiths Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111 A Short History of Research on Climate Change Impacts on Soil Microbial Communities. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 112 How can We Predict the Effect of Climate Change on Soil Microbial Communities? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 117 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123

6

Nitrous Oxide Production From Soils in the Future: Processes, Controls, and Responses to Climate Change . . . . . . . 131 Xia Zhu-Barker, Kerri L. Steenwerth Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 131 Biological Processes that Produce N2O in Soils . . . . . . . . . . . . . . . . . 133 Ammonia Oxidation Pathways . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 135 Heterotrophic Denitrification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 140 Other Biological Processes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 142 Abiotic N2O Production in Soils. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145 Hydroxylamine Decomposition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145 Chemodenitrification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 146 Land Management Practices to Control N2O Emission From Soils. . . . 147 Fertilization. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 147 Irrigation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 155 Tillage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 158 Cover Crops and Organic Amendments . . . . . . . . . . . . . . . . . . . . . . . 160

Contents Climate Change and Soil N2O Production. Conclusions. . . . . . . . . . . . . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . .

.. .. .. ..

. . . .

. . . .

. . . .

.. .. .. ..

. . . .

. . . .

. . . .

.. .. .. ..

. . . .

. . . .

. . . .

.. .. .. ..

. . . .

. . . .

. . . .

vii . . . .

162 164 165 165

7

The Response of Forest Ecosystems to Climate Change . . . . . . . 185 Armando Go´mez-Guerrero, Timothy Doane Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 185 Global Distribution of Studies on Climate Change and Forest Soils . . . 187 Changes in Net Primary Productivity of Forest Ecosystems . . . . . . . . . 188 Sequestration of Carbon in Forest Soils . . . . . . . . . . . . . . . . . . . . . . . 193 The Capacity of Forest Soils to Provide Ecosystem Services . . . . . . . . 195 Soil Processes in Relation to Soil Texture . . . . . . . . . . . . . . . . . . . . . . 196 Microbial Processes in Forest Soils . . . . . . . . . . . . . . . . . . . . . . . . . . 198 Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201

8

Effects of Elevated CO2 in the Atmosphere on Soil C and N Turnover . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 207 Yakov Kuzyakov, William R. Horwath, Maxim Dorodnikov, Evgenia Blagodatskaya Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 207 Approaches to Investigate Indirect Effects of Elevated CO2 Concentration on Soil Processes . . . . . . . . . . . . . . . . . . . . . . . . . . 208 Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 211 Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 216 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 217

Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 221

This page intentionally left blank

List of Contributors Evgenia Blagodatskaya Department of Soil Science of Temperate Ecosystems, Department of Agricultural Soil Science, University of Goëttingen, Goëttingen, Germany

Eric C. Brevik Department of Natural Sciences, Dickinson State University, Dickinson, ND, United States

Franciska T. de Vries School of Earth and Environmental Sciences, The University of Manchester, Manchester, United Kingdom

Timothy Doane Department of Land, Air and Water Resources, University of California, Davis, CA, United States

Maxim Dorodnikov Department of Soil Science of Temperate Ecosystems, Department of Agricultural Soil Science, University of Goëttingen, Goëttingen, Germany

Armando G omez-Guerrero Colegio de Postgraduados, Postgrado en Ciencias Forestales, Carretera Me´xico-Texcoco, Montecillo, Estado de Me´xico

Robert I. Griffiths Centre for Ecology and Hydrology, Wallingford, United Kingdom

Katherine Heckman Northern Research Station, USDA Forest Service, Houghton, MI, United States

Jeffrey A. Homburg Statistical Research, Inc., Tucson, AZ, United States; School of Anthropology, University of Arizona, Tucson, AZ, United States

William R. Horwath Department Land, Air and Water Resources, University of California, Davis, CA, United States

ix

x

List of Contributors

Yakov Kuzyakov Department of Soil Science of Temperate Ecosystems, Department of Agricultural Soil Science, University of Goëttingen, Goëttingen, Germany

Hans Lambers School of Plant Biology, Faculty of Natural and Agricultural Sciences, The University of Western Australia, Crawley, Australia

Craig Rasmussen Soil, Water & Environmental Science Department, University of Arizona, Tucson, AZ, United States

Jonathan A. Sandor Agronomy Department, Iowa State University, Ames, IA, United States

Lucas C.R. Silva Environmental Studies Program, Department of Geography, Institute of Ecology & Evolution, University of Oregon, Eugene, Oregon, United States

Kerri L. Steenwerth USDA-ARS, Crops Pathology and Genetics Research Unit, Department of Viticulture and Enology, University of California, Davis, CA, United States

Xia Zhu-Barker Department of Land, Air and Water Resource, University of California, Davis, CA, United States

Chapter | One

Soils, Climate, and Ancient Civilizations

Eric C. Brevik*, 1, Jeffrey A. Homburgx, {, Jonathan A. Sandorjj

*Department of Natural Sciences, Dickinson State University, Dickinson, ND, United States; xStatistical Research, Inc., Tucson, AZ, United States; { School of Anthropology, University of Arizona, Tucson, AZ, United States; jj Agronomy Department, Iowa State University, Ames, IA, United States 1

Corresponding author

INTRODUCTION Civilization as we know it today is closely tied to agriculture, as the adoption of agriculture and the more steady, reliable, and higher quantity of food it supplies promoted increased populations and the development of permanent settlements, growing populations beyond numbers that could be supported by hunting and gathering (Binford et al., 1997; Kirch, 2005; Montgomery, 2007). Agriculture in turn is tied to soil, as soil supplies several major requirements such as an anchoring medium as well as water and nutrient supply and storage that are necessary to propagate crops. Therefore, civilization is tied to soil. In the earliest days of agricultural production, human soil knowledge was rudimentary (Brevik and Hartemink, 2010), but before the end of the Neolithic settlements were being located at sites with rich soils well suited to agriculture (Montgomery, 2007; Miller and Schaetzl, 2014), indicating that human knowledge of the soil properties needed for good crop growth was developing. Just as human civilization relies on soils as the base of agriculture, it has also been highly influenced by our planet’s ever evolving climate. Fluctuations between relatively wet versus dry (Weiss et al., 1993; Binford et al., 1997) or warm versus cold (McMichael, 2003) climates have contributed to civilization rise and collapse throughout human history, as these climatic fluctuations influenced the ability of humans to produce agricultural products at levels needed to sustain them. However, it is also important to note that climate change is typically only one aspect that may contribute to civilization collapse. Other factors such as social and economic conditions influence the health of civilizations and Climate Change Impacts on Soil Processes and Ecosystem Properties. https://doi.org/10.1016/B978-0-444-63865-6.00001-6 Copyright © 2018 Elsevier B.V. All rights reserved.

1

2


the way a given civilization responds to climate change (Kirch, 2005). For example, clinging to cultural traditions has been suggested as an explanation for the failure of the Norse settlements in Greenland as the Little Ice Age brought on a colder climate, given that Inuit communities were able to survive the same climate change (Pringle, 1997). Climate change and its impact on both soils and civilizations is a major topic of interest today (Fig. 1.1). Studying what has happened in the past during changing climates can help us understand what is likely to happen in the future, and studying the way that past people have either adapted or failed to adapt to changes in climate can provide insight into potentially successful versus unsuccessful strategies to adapt to future climate change. Therefore, understanding our past is an important part of planning for the future.

FIGURE 1.1 Change in average surface temperature (A) and change in average precipitation (B) based on multimodel mean projections for 2081e2100 relative to 1986e2005 under the RCP2.6 (left) and RCP8.5 (right) scenarios. The number of models used to calculate the multimodel mean is indicated in the upper right corner of each panel. Stippling (i.e., dots) shows regions where the projected change is large compared to natural internal variability and where at least 90% of models agree on the sign of change. Hatching (i.e., diagonal lines) shows regions where the projected change is less than one standard deviation of the natural internal variability. Figure courtesy of IPCC, 2014. Climate change 2014: synthesis report. In: Core Writing Team, Pachauri, R.K., Meyer, L.A., (Eds.), Contribution of Working Groups I, II and III to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. IPCC, Geneva.

The Use of Soils in Archaeology

3

THE USE OF SOILS IN ARCHAEOLOGY Understanding ancient civilizations is the realm of archaeology, and soil science can be an important tool used by archaeologists to supply information on a wide range of subjects (Holliday, 2004; Homburg, 2005). Soils are an integral component of cultural landscapes and can provide significant information to archeological studies. This may include the impact of human occupation on a site and the environmental setting at the time of human occupation (Holliday, 2004). The location of buried soils can be used as markers for where artifacts are likely to be found (Hardman et al., 1998), and the location of artifacts within a soil can sometimes be used to assign approximate dates to the artifacts (Homburg, 1988). The number of soils at a site and the degree to which each soil developed can also provide important information about how much time a given archaeological site spans, the integrity of its archaeological record (Holliday, 1992), landscape evolution, and environmental change over time (Jacob, 1995; Monger, 1995; Mayer et al., 2005; Homburg et al., 2014a). Therefore, beyond just being an important part of the story about ancient civilizations, soils can also be an integral part of deciphering and understanding those civilizations. Soils have been useful to archaeology in the study of ancient agricultural systems, as they provide significant insight into the diet (Sweetwood et al., 2009) and general land use (Jacob, 1995; Homburg and Sandor, 2011) of ancient people. Soil chemistry has been applied to help interpret activity patterns of occupation surfaces within a site (Entwistle, 2000; Parnell et al., 2002; Macphail et al., 2004). Soil micromorphological analysis has been used to reconstruct past soil management practices (Wilson et al., 2002) and activity areas within a site (Macphail et al., 2004) and interpret the intensity of various economic activities (Simpson et al., 2005) (Fig. 1.2). Induced magnetism in soil is a well-established technique used by archaeologists to locate, characterize (Smekalova et al., 1993; Schmidt, 2007), date (Schmidt, 2007; Hambach et al., 2008), and interpret archaeological sites, features, and stratigraphy. Soils have frequently been used as part of broader, interdisciplinary approaches in archaeology. Sarris et al. (2004) used a combination of geophysics and soil chemistry to reconstruct a Copper Age settlement in modern-day Hungary. Homburg et al. (2014b) used sedimentology, stratigraphy, macrofossil and microfossil analysis, radiocarbon dating, and soils to reconstruct 7000 years of lagoon evolution in southern California, USA, to direct archaeological investigations of areas within a large study site that were most likely to yield significant cultural deposits, and May et al. (2008) used a combination of soils, stratigraphy, and artifacts to reconstruct the paleoecology of archaeological sites in Nebraska, USA. When used in combination with other fields, soils provide a very powerful tool in paleoenvironmental reconstruction, which is valuable to archaeologists who are attempting to decipher the resources that were available to prehistoric humans (Homburg, 2005). These studies can also be useful in the reconstruction of past climate changes and understanding how those changes may have influenced ancient civilizations.

4


FIGURE 1.2 Soil samples being collected for eventual thin section analysis (rectangular areas in the profile) at an archaeological excavation. Artifacts are also visible in the photograph. Figure courtesy of Jeffrey Homburg.

STUDIES AT ANCIENT SITES TO UNDERSTAND SOILS Deep time is a concept of geologic time first developed in the 1700s by the Scottish geologist, James Hutton (1726e97), whereby he envisioned the geologic record as the product of endless slow cycles of rock formation below sea, including uplift and tilting of rock layers, subsequent erosion, and then formation of new strata below the sea. Boucher de Perthes (1847) was the first to demonstrate the deep antiquity of humans when he recovered ancient stone tools in stratigraphic association with bones of extinct Pleistocene mammoths, cave bears, and other mammals at a cave site in the Somme valley of northern France. This study demonstrated that humanity spans much longer than the 6000 years of time that was then accepted, following Archbishop James Ussher’s (1581e1656) opinion more than two centuries earlier (Grayson, 1983). Similar to the geologic record, cycles of change can also be identified in the archaeological record, in this case over the life spans of civilizations.

Studies at Ancient Sites to Understand Soils

5

Cycles of flooding, drought, earthquakes, volcanic eruptions, soil erosion, and other kinds of environmental hazards, as well as the spread of disease or warfare, have been very disruptive to past human civilizations, sometimes enough to explain their collapse. Archaeology, like geology and soil science, is a historical science, and just as soils can provide insights into archaeology; archaeology can in turn provide insights into soils. The archaeological record takes a significant time to form, as cultural artifacts and other materials (e.g., pollen, phytoliths, charred plant remains, animal bone, etc.) are incorporated and embedded within the sedimentary and soil matrix of an archaeological site. Physical and chemical soil properties are commonly used by geoarchaeologists to make interpretations of the archaeological record, including how it was formed and altered in the past and how it is preserved today (Schiffer, 1987). Soils are used to draw archaeological inferences about human behavior, including identifying chemical signatures of different kinds of human activities in the past that may leave no physical traces. Archaeological studies are also used for assessing the long-term effects of cultivation on soil productivity. The rest of this section focuses on methods and uncertainties in evaluating soil productivity in the archaeological record. Soil properties used for identifying evidence of agricultural degradation are summarized in Table 1.1. Use of the archaeological record has certain limitations and advantages and these are reviewed in the following section.

Limitations The archaeological record is an imperfect record that is affected by numerous cultural and natural factors. In addition, there are a number of potential methodological problems and uncertainties that affect evaluations of the long-term effects of agriculture on soils in the archaeological record (Sandor and Homburg, 2011). Among these are difficulties in (1) field identification, (2) variability in the kind and age of agricultural systems, and their impact on soil, (3) postagricultural environmental change and land use impacts on soils, (4) availability of appropriate unfarmed (“control”) soils to use as references for quantifying soil change from agriculture; kinds of control soils, their validity, and what can be inferred from them, (5) sample designdnumber, depth, and type of samples and sample sites needed to test for soil differences, and (6) appropriate physical, chemical, and biological assays of soil properties and how to interpret results (e.g., Holliday, 2004; Homburg et al., 2005; Sandor et al., 1986). The potential and pitfalls of the comparative approach and other soil change testing issues are reviewed in more detail by Homburg and Sandor (2011).

Advantages and Benefits The main advantages of the archaeological record are that it provides a deep time perspective for testing ideas and hypotheses, ideas that sometimes are

6


Table 1.1 Soil Properties that May Indicate Soil Degradation Caused by Cultivation Soil Property

Criteria for Recognizing Degradation: Typical Causes and Consequences

A horizon thickness

Decreased thickness caused by water or wind erosion. Reduces important organic mattereenriched surface layer that can be exploited by plants for water, nutrients, and oxygen. Shallower depth to possible root-limiting subsurface layers such as strongly developed argillic horizons.

Soil structure

Macromorphology: lowered grade of granular or subangular blocky structure, trend toward massive state, especially in surface horizons. Commonly results from compaction and organic matter decline.

Bulk density

Compaction (increase in bulk density above that of natural condition) associated with soil structure degradation. Compaction and structure degradation commonly retard seed germination and root growth, reduce root access to water, oxygen, and nutrients, reduce diffusion of gases, and decrease water infiltration and available water capacity.

Organic carbon

Decrease in organic C is common under conventional cultivation. Results from accelerated microbial oxidation of organic matter in disrupted, exposed soil aggregates, and other effects of agriculture. Numerous benefits of organic matter for soil physical, chemical, and biological properties important to plant growth are well documented.

Nitrogen

Decrease in total N accompanies declining organic matter in agricultural soils, although C:N ratio tends to decrease. Nitrate and ammonium are plant available forms of N, which is commonly a key limiting factor for plant growth in all regions, including arid regions.

Phosphorus

P (both total and available) is another macronutrient that has been shown to decrease due to cultivation in some cases. P is a key ecological and soil indicator because of its low mobility, low availability to plants, and long-term stability of its forms in soils.

pH

Very high soil pH can indicate salt accumulation (which is measured by electrical conductivity). Sodic soil conditions (recognized by high exchangeable sodium) can be prevalent in agricultural soils of arid and semiarid regions. Detrimental effects on many plants, including crop species, occur both through direct chemical effects and through soil structural deterioration. pH is also an indicator of the availability of nutrients to crops.

Adapted from Homburg et al., (2005).

derived from historical accounts and documentary evidence. However, much of the archaeological record predates historical accounts, often by centuries to millennia, so history is mute for much of the deep time record of archaeology. Neither the archaeological nor historical sources are complete records, but they can sometimes be combined and integrated (especially at that nexus between the two records), thereby providing a powerful method for testing the other discipline. Archaeology aims to reveal information about the past by excavating artifacts, cultural features (e.g., houses, storage facilities, ceramic kilns, but also canals, terraces, fields, and other kinds of agricultural features), and associated environmental samples such as soil and pollen samples. These samples are then

Soil Knowledge and Management in Early Civilizations

7

analyzed and used for reconstructing cultural activities and assessing human effects on the environment (e.g., the timing of forest clearance for agriculture, effects of agriculture on soil health).

SOIL KNOWLEDGE AND MANAGEMENT IN EARLY CIVILIZATIONS Many early civilizations demonstrated knowledge of soil management as a way to improve agricultural production, and much of our understanding of this early soil knowledge comes from archaeological studies. The earliest evidence of soil manipulation for agricultural production dates back to approximately 9000 BCE (Troeh et al., 2004). Between 7500 and 500 BCE a number of other innovations were developed, including irrigation (Troeh et al., 2004), terracing (Sandor, 2006), early plows, and contour tillage (Brevik and Hartemink, 2010). While a trial and error approach was most likely used by early farmers to find the best agricultural sites, evidence indicates that soil spatial patterns were being recognized by some civilizations and used to select cropping (Krupenikov, 1992) and settlement (Miller and Schaetzl, 2014) sites by 3000e2000 BCE. Early civilizations such as the Chinese (Li and Cao, 1990; Gong et al., 2003) and Greeks (Krupenikov, 1992) were also classifying soils by 300 BCE. The study of soils had not yet advanced to the point of being a science in these early civilizations; soil science becoming a modern scientific field would not happen until approximately 1883 CE (Coffey, 1911; Landa and Brevik, 2015). However, ancient human populations around the world were definitely aware of soils and ways to manipulate those soils to their advantage, and Krupenikov (1992) argues that soil science appeared as an empirical field about 500e0 BCE. Some examples of this early knowledge from different parts of the world will be reviewed in this section.

Asia The oldest indications of human management of soil found to date come from Asia, near Jarmo in Iraq, and date to about 9000 BCE (Troeh et al., 2004; Montgomery, 2007). Evidence of agricultural production showed up shortly afterward in Abu Hureyra in modern Syria (Montgomery, 2007). Irrigation was developed in modern-day Iraq by 7500 BCE (Troeh et al., 2004), and early plows were developed in the Middle East between 6000 and 4000 BCE, which revolutionized soil preparation for planting (Hillel, 1991; Lal, 2007). The plow appeared about the same time that all fertile land in Mesopotamia had been placed under cultivation, and it was necessary to find ways to increase crop production by means other than simply adding new agricultural fields (Montgomery, 2007). With the domestication of sheep and goats about 8000 BCE in the Zagros Mountains area, manure was used to fertilize fields (Montgomery, 2007), and occupants of the marshes in southern Iraq were creating artificial soils to build up small islands in the marshes and support their agricultural lifestyle as early

8


as 3000 BCE (Fitzpatrick, 2004). Terraces were utilized for the first time in the Near East between 4000 and 3000 BCE (Sandor, 2006). Ancient civilizations in other parts of Asia were also developing agriculture and a knowledge of soil. Agricultural production began in China not long after its beginnings in the Middle East (Montgomery, 2007). Other ancient indications of soil management for agricultural purposes come from China, where rice was cultivated by 7000 BCE (Gong et al., 2003), paddy soils were in use by 4000 BCE (Cao et al., 2006), and terraces by 1000 BCE (Sandor, 2006) (Fig. 1.3); India, where plowing was a common practice by 3000 BCE (Brevik and Hartemink, 2010); and Uzbekistan, where farmers amended soils to improve both fertility and texture as early as 2000 BCE (Krupenikov, 1992). Early drainage canals found in Papua New Guinea date to 7000 BCE with more extensive drainage systems and the construction of raised soil beds by 4000 BCE (Vasey, 2002).

Europe Agriculture spread west from the Middle East into Europe, starting with Turkey and Greece about 6300 BCE (Montgomery, 2007). The ancient Greek philosopher-scientists recognized differences between soils by the second millennium BCE; they are credited with the first recorded works that show knowledge of soil properties and with developing a soil profile concept (Krupenikov, 1992). They understood that plant nutrients were supplied by soil (Sparks, 2006) and wrote of water storage in soils (Hillel, 1991). Their attention to soils allowed the Greeks to choose crops best suited to the soils at a given location (Krupenikov, 1992). Despite this, soil erosion and degradation became a major problem in ancient Greece (Dotterweich, 2013; Hillel, 1991; Troeh et al., 2004). The Romans utilized manure and green manure (Fig. 1.4) to enhance soil fertility (Winiwarter, 2006) and used terraces to reduce erosion (Troeh et al., 2004). The Romans also had a system of soil classification and developed tests for soil fertility (Tisdale et al., 1993; Winiwarter, 2006). Because the Roman Empire encircled the Mediterranean and reached far into western and northern Europe, Roman ideas on soils and agriculture were widely influential. In other

FIGURE 1.3 Rice terraces in Banaue, Ifugao, the Philippines (left) and farmers working in a terraced rice paddy in the Philippines (right). Photographs courtesy of IRRI Photos/Gene Hettel.


9

FIGURE 1.4 Cover crops, which can be used as green manure, growing in an orchard in California. Photograph courtesy of Gary Kramer, USDA-NRCS.

parts of Europe, contour tillage was developed in the British Isles prior to the Roman conquest in 43 CE (Troeh et al., 2004), and some cultivated land on steep slopes was reforested as a soil conservation measure in parts of Europe as early as 900 CE (Krupenikov, 1992).

Africa North Africa had some very successful ancient agricultural societies. The ancient Egyptian civilization lasted for 3000 years, from 3300 BCE to 332 BCE, based on an agricultural system where fertility was sustained through frequent flooding of the Nile River and crop water needs were met through irrigation using water from the Nile (Hillel, 1991; Troeh et al., 2004). The Egyptians utilized cultivation to prepare the soil for sowing (Krupenikov, 1992). Farther west, Carthage was founded between 835 and 800 BCE (Aubet, 2008) and became a political power about 600 BCE, with a solid agricultural

10


base that produced cereals, fruits, vegetables, and nuts (Giamrnellaro, 1999). The Roman conquest of Carthage in 146 BCE was in part to secure land that could produce the crops necessary to feed Rome, and the Romans were very interested in Mago’s handbook of Carthaginian agriculture (Montgomery, 2007). The Roman agricultural writer Lucius Junius Moderatus Columella referred to Mago as “The Father of Agriculture” (Mahaffy, 1889). The Carthaginians utilized advanced cultivation and irrigation systems. However, wind and water erosion eventually removed the topsoil around Carthage, leaving the region incapable of supporting the populations it once did (Troeh et al., 2004).

Americas The earliest evidence of agriculture in the Americas dates to between 8500 and 7000 BCE in Peru and Mexico (Vasey, 2002). Maize was domesticated by 7000 BCE (Larson et al., 2014), which led small dispersed villages in Mesoamerica to grow into larger towns that served as cultural and economic centers (Montgomery, 2007). From 600 BCE until about 800e900 CE the Mayan civilization thrived in Central America, with a population that increased from only around 200,000 at its beginning to 3e6 million at its height (Montgomery, 2007). This population was supported by a dynamic and ever evolving agricultural system that differed significantly both geographically within the Mayan empire and temporally as the population increased; agricultural intensification tended to increase as the population grew (Dunning et al., 1998). The Maya progressed from a system of slash and burn agriculture under low population density to more intensive raised field systems with extensive drainage canal networks in lowland locations and extensive terraces on steeper slopes, but a lack of domesticated animals meant that Mayan agriculture lacked manure for renewing soil fertility. Other population centers in Central America and Mexico had similar agricultural adaptations (Montgomery, 2007). Cultivation of crops by some Native American tribes began in the Midwestern and Northeastern USA by 5000 BCE (Warren, 1994). In this area intercropping diverse crop mixes, additions of ash, and fallowing were used to maintain soil fertility, and crops were planted in small mounds that were more resistant to erosion than row cropping (Brevik et al., 2016a) (Fig. 1.5). Ancient agricultural management strategies employed in the American Southwest, such as runoff terraces, rock mulch, and irrigation, were matched to soil and landscape settings (Homburg and Sandor, 2011). Domesticated maize diffused to North America from Mesoamerica, appearing in the crops of North American societies by approximately 2000 BCE (Hart et al., 2007; Drake et al., 2012; Huber, 2004). Agricultural terraces date back at least 4000 years in Peru, and many ancient terraces are still farmed today (Denevan, 2001; Sandor, 2006; Sandor and Eash, 1995) (Fig. 1.6). Intercropping, crop rotations that included legumes, fallowing, manuring, and the addition of ash were used to maintain soil fertility (Montgomery, 2007). Animal domestication in South America (Larson et al., 2014) provided a source of manure that was absent in other parts of the


11

FIGURE 1.5 Intercropped maize and squash in a traditional Wampanoag garden in Massachusetts, USA. The maize is planted in small mounds, most noticeable in the lower right corner of this figure. Photograph courtesy of Eric Brevik.

FIGURE 1.6 Ancient agricultural terraces in the Colca Valley, Peru. Terraces still farmed on lower slopes, abandoned terraces visible on upper slopes. Photograph courtesy of Jonathan Sandor.

Americas (Montgomery, 2007). In the Amazon, human activities created small patches of Dark Earth, or terra preta, soils between approximately 500 BCE and 1640 CE (Heckenberger et al., 2003; Neves et al., 2003). The thick A horizons and high organic carbon and nutrient content of these soils compared to surrounding natural soils promoted sustainable agricultural production in the

12


FIGURE 1.7 An area of terra preta (left) compared to a nearby natural area with an Oxisol (right) in the central Amazon Basin of Brazil. Note the much deeper and darker A horizon of the terra preta, indicating the impact of anthropogenic activity on this soil. Photographs courtesy of David Laird.

terra preta (Fig. 1.7), but it is not currently known if the terra preta was created intentionally or as an accidental by-product of human activities (Glaser and Birk, 2014). Native Americans in the Bolivian Amazon established raised fields and canal systems that supported complex societies with a relatively high population density (Lombardo et al., 2015).

EFFECTS OF ANCIENT AGRICULTURE ON SOILS AND SOCIETIES Given that humans have been manipulating soil environments for at least 11,000 years to improve agricultural production, humans have also influenced soil properties and processes over the same time span. While earlier humans lacked the ability to manipulate as much soil as quickly as we can today, the effects of their work can still be seen in many places. We often focus on degradation issues when discussing humans and our effects on the soil resource (Sandor et al., 2005; Richter, 2007; Fernańdez-Romero et al., 2014; Ibań˜ez et al., 2015; Khaledian et al., 2017), but there are examples of societies who developed sustainable practices as well. We can learn from both the sustainable and unsustainable choices that earlier civilizations made. This section will provide some examples of both sustainable ancient systems and situations where ancient agriculture may have caused severe soil degradation, enough to have either caused or at least contributed to civilization decline, with some discussion of what we can learn from each example.

Effects of Ancient Agriculture on Soils and Societies 13

Examples of Sustainability While Mesopotamia is often cited as a prime example of humans causing widespread land degradation that contributed to eventual civilization collapse (Yaalon and Arnold, 2000; Rengasamy, 2006; Lal, 2009; Brevik and Hartemink, 2010), an agriculturally based culture developed in the marshlands along the southern reaches of the Tigris and Euphrates rivers that continued for approximately 5000 years (Fitzpatrick, 2013). These marsh dwellers practiced a low-intensity agricultural lifestyle from about 3000 BCE until 1980 CE that consisted of building up small islands to support their reed houses and to raise dates, vegetables, and water buffalo (Buringh, 1960; Fitzpatrick, 2004). High concentrations of N and P in the soils of these small islands indicate a slow buildup of organic-rich materials through anthropogenic activity over long time periods (Fitzpatrick, 2004). Subsidence of the Earth’s crust meant that the marshes remained at a low elevation with wet conditions even as the Tigris and Euphrates rivers deposited sediments in the area; this process also required constant buildup of the marsh islands by their occupants (Buringh, 1960). This lifestyle proved to be sustainable until large-scale dam building occurred along the Tigris and Euphrates and marsh drainage programs were implemented in the 1970s and 1980s (Fitzpatrick, 2004). The resulting change in soilewater relationships lead to the collapse of a lifestyle that had been sustainable over thousands of years. Future soilewater relationships are also expected to change in at least some ecosystems under climate change (Brevik, 2012). The lesson learned from this example is that it is possible to find ways to coexist harmoniously with at least some natural systems such that very long-term (by human standards) sustainable lifestyles are possible, but changing environmental conditions can upset that balance. In his classic work “Farmers of Forty Centuries,” King (1911) recognized that despite dense populations with a low area of arable land per person, civilizations in eastern Asia had managed to maintain sustainable agricultural production for as long as 4000 years. King attributed the success of the Asian farmers to several factors, including (1) a favorable geographic location with abundant rain and growing seasons long enough to allow multiple crops per year, sometimes as many as four, particularly with the use of intercropping, (2) good crop selection for local conditions, (3) high appreciation for the value of water in crop production leading to good water management through the combination of irrigated and dryland cropping systems on the same land, depending on season, (4) use of mulches for water conservation during drier periods, and (5) use of multiple nutrient sources as fertilizers, including green manure, composts, ashes, animal and human manure, and growing legumes in crop rotations. King was particularly impressed with the fact that nothing organic went to waste; it was either consumed by humans or animals, worn as clothing, used as fuel, or returned to farm fields as fertilizer. These practices maintained soil fertility and productivity by effectively cycling nutrients and replenishing soil organic matter. The lesson learned is that even intense

14


agricultural systems can be sustainable if managed properly. It is noteworthy that King concluded that favorable climate was one essential ingredient that allowed East Asian agriculture as it was practiced to be sustainable. The obvious corollary is that climate change could create a situation within which the same practices were no longer sustainable. There are agricultural soils in the Andes Mountains of South America that have been cultivated for at least 1500 years and still remain productive (Sandor and Eash, 1995). Although located in steep mountainous terrain, building terraces created gentle slopes with arable land and a way to manage irrigation. Manure and ash from the farmers’ hearths were incorporated into the terraced soils as nutrient sources. These practices created A horizons in terraced soils that had lower bulk density, higher C, N, and P levels, and thicker A horizons than nearby, natural uncultivated soils on steep mountain slopes (Sandor and Eash, 1995). After 1500 years of agricultural production in the Colca Valley, Peru, locals continue to farm terraces sustainably. In addition, rehabilitation and restoration of some abandoned ancient Andean terraces in recent years has led to increased crop production and financial gain for farmers who have participated in trials (Branch et al., 2007). Some researchers have questioned the sustainability of these terrace systems during climatic fluctuations that led to prolonged dry periods (Branch et al., 2007), showing one possible climate/agricultural management interaction that may have influenced ancient cultures and the sustainability of their agricultural practices. One of the major lessons learned here is that a given system may be sustainable under some climates, but not others.

Examples of Degradation Unfortunately, it is not difficult to find examples of soil degradation in the annals of human history. One of the most frequently cited examples of civilization failure due to soil degradation involves the civilizations of Mesopotamia. The Sumerians saw their political power in Mesopotamia wane as their soils became too saline and waterlogged for crop growth because of salts in irrigation water and rising groundwater levels caused by irrigation (Hillel, 1991). The Babylonians became the dominant political power in Mesopotamia after the Sumerians until the Babylonian canals filled with silt eroded off the surrounding hills, caused by increased erosion rates triggered by clearing forests from hills that bordered the river valleys to provide timber and create grazing areas for their livestock (Troeh et al., 2004). Along the borders of southern Iraq’s marshlands, the same location where the marsh dwellers successfully used the same agricultural practices for 5000 years, areas were utilized for rice production as early as 550 BCE. These areas experienced accelerated sedimentation that filled in the rice fields over time. After a rice field was no longer useful due to sediment buildup, the field would be abandoned and new rice fields created, leaving behind an area of largely unproductive land (Buringh, 1960). The lesson learned from the Mesopotamian experience is that proper water management and soil conservation practices are critical to sustainable agriculture.

Effects of Ancient Agriculture on Soils and Societies 15 Another classic example of cultural collapse due to soil degradation and resource exploitation is found on Easter Island. Polynesian settlers established an agricultural society on Easter Island (Rapa Nui) approximately 1200 CE (Hunt and Lipo, 2006; Mann et al., 2008), with deforestation and extensive soil erosion beginning within the first century of occupation (Hunt, 2007; Mann et al., 2008). These early settlers arrived on an island vegetated with a mesic forest including Paschalococos disperta palm (Mann et al., 2008), but deforestation was largely complete by 1650 CE (Hunt, 2007). At one time Easter Island may have supported a population as high as 10,000 inhabitants (Diamond, 2005); however, by 1650 CE the society collapsed and the population declined steeply (Mann et al., 2008). When he visited in 1774 James Cook found only about 1000 people living on Easter Island (Heyerdahl, 1961). Three main hypotheses have been offered to explain the collapse of the Polynesian society on Easter Island: (1) overexploitation and degradation of the island’s resources, especially palm trees, led to a situation where the island could no longer support its inhabitants, (2) contact with European explorers which led to problems such as disease outbreaks, and (3) environmental changes beyond human control triggered civilization collapse (Mann et al., 2008). Of these, the first currently seems to be the leading explanation. If the first hypothesis is true, Easter Island serves as a warning about what can happen if humans abuse the ecosystems that support them, and if the third hypothesis is true it serves as a warning regarding the potential impact of environmental changes, such as climate change, on human societies. It is also possible that a combination of these is true, e.g., degradation of the environment created a weakened society that was ill equipped to survive an environmental change. This is a sobering possibility as we regard a modern world with an environment that is degraded in many ways coupled with a changing climate. While Mesopotamia and Easter Island are among the most commonly cited examples of anthropogenically driven environmental degradation leading to civilization collapse, alternative interpretations to the evidence found in these locations has been offered. For example, Hunt and Lipo (2006) and Hunt (2007) have argued that the loss of trees on Easter Island was due to the invasion of rats (Rattus exulans) that travelled to Easter Island with the human settlers, that these rats multiplied rapidly and devastated the island’s trees by eating seeds, roots, and tree sprouts, and that there is no solid evidence that the earliest settlers used slash and burn management. However, Mann et al. (2008) dispute the Hunt and Lipo hypothesis, stating that lake cores from Easter Island have strong evidence of widespread burning and soil erosion coinciding with human settlement and that they could find no evidence for a “rat outbreak impact.” Likewise, Powell (1985) and Rost (2015) have questioned the interpretation that soil salinization was a contributing factor to the collapse of Mesopotamian civilizations. Examples such as these demonstrate some of the difficulties and ambiguities in interpreting the archaeological record; the evidence available often has defendable alternative interpretations, even in situations like

16


Mesopotamia and Easter Island that are frequently invoked as examples of civilization collapse due to environmental degradation. Some civilization failures were likely caused by a combination of climate changes and cultural factors. A prime example was the Norse in Greenland. There is some debate over whether or not soil degradation was a problem for the Greenland settlements, with some researchers concluding there is no evidence of significant widespread soil degradation (Rutherford, 1995), whereas others have concluded erosion was a significant problem in at least some locations (Berglund, 1986). There are, for example, buildings at some sites that were buried by drifting sand, likely initiated by overgrazing (Berglund, 1986). However, there is little debate that climate at least played a role in the Norse demise. The climate in Greenland was unusually mild when the Norse first settled in Greenland in the late 900s CE, followed by the onset of the Little Ice Age about 1300 CE, and complete abandonment of Norse settlements in Greenland in the 1400s CE (Pringle, 1997). This led many researchers to conclude that the Norse settlements in Greenland failed due to climate change (Buckland et al., 1996; Pringle, 1997; Barlow et al., 1997). Studies have indicated a change in diet from primarily terrestrial-based food sources to greater dependency on marine-based food sources over the course of the Norse presence in Greenland; a cooling climate is one reason that may have spurred this dietary shift (Berglund, 1986; Arneborg et al., 1999). Despite adjustments such as a shift in diet, the Norse settlements in Greenland eventually failed even as Inuit communities survived the same climate change (Pringle, 1997). Berglund (1986) and Barlow et al. (1997) concluded that the decline of the Norse Greenland settlements was likely due to a combination of climate change and a conservative culture that was resistant to change, leaving the Norse vulnerable to environmental alterations. The takeaway message for the future is that land degradation is not the only challenge to the health of a civilization; cultural inflexibility may also be a serious liability in a world of changing climate.

Summary of Lessons to Be Learned Sustainable agricultural systems have been developed by people living in widely dispersed locations and under a variety of climate, landscape, and social conditions. King (1911) noted that eastern Asian cultures had successfully sustained themselves for 40 centuries due to the development of management practices that took full advantage of their environment. King (1911) argued that other societies, including industrialized western societies, had much that they could learn from the East Asian example, but he also acknowledged that other societies would not be able to simply copy the East Asian techniques because they lived under different environmental and social conditions. In today’s world we have numerous examples of scientifically developed agricultural systems created in industrialized countries that failed to work when they were introduced in developing countries, sometimes for cultural reasons (Critchley et al., 1994; Hellin and Haigh, 2002). In recent years this has led to increased

Climate Change and Ancient Cultures

17

FIGURE 1.8 A view looking along the Mormon Trail in Iowa, USA, August, 1999 (left), approximately 146 years after the trail was abandoned. Note that soil compaction has led to reduced forage growth in the trail. Soil organic carbon (SOC) content of the top 20 cm of the on trail and off trail soils (right). An analysis by 5 cm increments indicated that there were no statistically significant differences between SOC content in the top 10 cm of on trail versus off trail soils, but that the off trail SOC contents were significantly higher in the 10e20 cm depth intervals. Data from Brevik, E.C., Fenton, T.E., 2012. Long-term effects of compaction on soil properties along the Mormon Trail, south-central Iowa, USA. Soil Horizons 53 (5), 37e42. Photograph courtesy of Eric Brevik.

calls to incorporate local farmers’ knowledge into agricultural management planning to create sustainable systems (Barrera-Bassols and Toledo, 2005; Sandor et al., 2006; Mairura et al., 2008). Approximately 100 years after King’s seminal work, we are coming back to the recognition that there is not a “one size fits all” solution to sustainable agricultural management and that indigenous approaches that have been used for a long time may in fact have valuable lessons even for the scientific sector. Anthropogenic interactions have the ability to significantly alter the soil system. Soil disturbances by humans tend to increase microbial conversion of soil organic matter to carbon dioxide as well as the rates of soil erosion and nutrient loss, leading to soil degradation (Amundson et al., 2015). Intense trafficking can cause soil compaction and related problems such as reduced vegetative growth and soil organic matter additions (Brevik and Fenton, 2012) (Fig. 1.8), compaction that can sometimes persist for centuries (Sandor and Eash, 1991). To prevent negative anthropogenic impacts on the soil resource, it is important that we carefully study the experiences of past civilizations, both successes and failures, and learn from what they did.

CLIMATE CHANGE AND ANCIENT CULTURES Climate and climate change have influenced human societies throughout history. Climate amelioration at the end of the Pleistocene facilitated the beginnings of complex societies and agriculture. Many early civilizations in both the Old World (Mesopotamia, Egypt, Indus) and the Americas (Peru, highland Mexico, and American Southwest) arose in arid regions (Scarborough, 2003). Favorable climates have provided conditions for past societies to take root and flourish. On the other hand, climate change, especially prolonged drought

18


and its environmental effects, has been implicated as a factor in the demise and collapse of several past civilizations (de Menocal, 2001; Fagan, 2008; Weiss and Bradley, 2001). Societies may not only be negatively affected by climate change in an absolute sense but also by increased climate unpredictability (Ingram, 2015). Previous climate change during the Holocene that has affected societies globally (e.g., the Medieval Warm Period and the Little Ice Age) or regionally are attributed to ocean and atmospheric circulation cycles such as the El NinoeSouthern Oscillation, cyclic and episodic changes in solar radiation, and volcanic activity. Humans are now added to this list of climate change drivers. Although climate change has been a significant factor in social change past and present, it is also important to recognize that causes for the rise and fall of societies involve several complex interacting social, economic, and environmental variables. Views that climate change determined past social change are too simplistic because people are not passive in the face of environment and its variability, but actively respond and adapt (Kirch, 2005). Assigning causality in past social transformation or collapse is controversial. This complexity in the interaction of humans and environment is evident in the wide range of response of peoples to climate change over space and time. Some environments are inherently more risky than others, and some societies have been more resilient than others. Archaeologists distinguish different degrees of vulnerability of societies to instability and failure in the face of environmental change and other pressures (Nelson et al., 2012; Ingram, 2015). Some societies lacking the flexibility to respond to environmental change have been caught in a “rigidity trap.” It is also important to recognize the challenges in measuring and interpreting past climate change in studies of ancient societies (Ingram, 2015). There are several tools for reconstructing paleoclimate (Huckleberry, 2015; Fagan, 2008) with varying degrees of accuracy and precision, and multiple approaches are important. Even the finest of these climate markers, such as tree rings, which provide precipitation and temperature data on an annual scale, are not yet precise enough to reconstruct conditions during crop growth such as for stages of maize development (Adams, 2015). Cases from several world regions, climates, and cultures illustrate the broad range of climatic and environmental impacts on water resources, crop growing season conditions, and food production encountered by past societies. These include many combinations of environmental and social conditions and outcomes for past societies. Some researchers have identified collapses in ancient societies in which severe drought played a significant role. These have occurred in northern China, the Indus Valley, Mesopotamia, Peru and the Andes, the Maya region, and North America (Benson and Berry, 2009; Benson et al., 2009; Binford et al., 1997; de Menocal, 2001; Doyel and Dean, 2006; Dunning et al., 2012; Fagan, 2008; Hoggarth et al., 2016; Meeks and Anderson, 2013; Montgomery, 2007; Moseley, 2001; Scarborough, 2003; Schwindt et al., 2016; Weiss and Bradley, 2001). Societal collapse has involved political upheaval, conflict, and long-term depopulation. To illustrate the complexity of

Climate Change and Ancient Cultures

19

social transformation in relation to climate change, the same drought of the mid-1100s AD in the American Southwest associated with the demise and depopulation of Chaco Canyon, affected other Southwestern cultures in the Mesa Verde and Mimbres areas (local population shifts) differently (Ingram, 2015). In some cases, systematic demographic and land use change in response to climate change has been a consistent component of some social structures, as in the “verticality” adjustments of Andean people and agriculture to higher moister elevations during drier periods or to lower elevations during wetter periods (Moseley, 2001). There are a number of cases in which ancient and traditional societies were able to respond successfully to climate change and remain in place. Of potential relevance today are the innovations developed by past peoples in actively adapting to climate challenges and change, including new technologies, crops, and agricultural strategies. For example, varieties of maize were developed in the Americas that were adapted to arid conditions or high altitudes (Adams, 2015). In the American Southwest, desert plants such as agave were domesticated in part to offset food shortages when other crops failed. A number of these adapted crops, the landscapes on which they were grown, and the techniques by which they were managed, are not used today but have potential (Minnis, 2014; Nabhan, 2013; Sandor and Homburg, 2015). One of the most important principles in adaptive traditional agricultural systems is their diversity in terms of crops and field location. Agricultural diversification is a deliberate strategy of risk management. Strategies of landscape risk management for climate variability, widely practiced in arid regions such as the American Southwest, include spatially diverse field locations, combinations of irrigated agriculture in valleys with dryland farming at valley edges and in uplands, and specially adapted crops and crop management, combined with food storage techniques. Irrigated agriculture may have been the mainstay, but upland production could provide a hedge against damaging floods or freezes resulting from cold air drainage into valley floors. Other examples of innovative ancient agricultural technologies developed in response to climate fluctuation, water scarcity, and climate changeerelated hazards such as erosion of agricultural terraces, lithic mulch, and other rock configuration fields for moisture conservation and temperature moderation, water harvesting, wells, and qanats (a gently sloping underground channel with vertical shafts used to access and transport water from groundwater to agricultural fields) (Adams, 2015; Doolittle, 2000; Hamidian et al., 2015; Nabhan, 2013; Sandor, 2006; Scarborough, 2003). Climate change not only impacts societies directly but also through complex interactive effects among hydrologic, geomorphic, and soil processes, and vegetation change (Bull, 1991). These changes occur on low- to highfrequency time scales and across spatial scales (Dean, 1988a,b; Huckleberry, 2015; Waters, 2006). Positive or negative outcomes in terms of impact on agriculture and society are possible, but negative outcomes are more likely. For example, stream instability involving channel widening and incision after centuries of stability, brought about in part by climate fluctuation, is thought to

20


have played a role in the social reorganization of prehistoric Hohokam society through disruption of their extensive canal irrigation systems (Waters, 2006). Another key point is that it can be difficult to separate the causes of land degradation from either natural climate change or anthropogenic impacts; they often operate together through feedback processes. Degradational effects of climate change may be triggered by anthropogenic landscape and ecological change or vice versa. There are instances in which climate change may not have caused landscape degradation such as accelerated erosion had it not been set up by anthropogenic change. An added dimension to this in the present day is human-induced climate change. Researchers infer a linkage of climate change and anthropogenic environmental change that has negatively impacted many past societies. Soil geomorphic and ecological degradation associated with climate change and agriculture has been documented in several world regions. Common examples include accelerated soil erosion in sloping agricultural lands in China, Europe, the Mediterranean and Middle East, the Mayan area, and the American Southwest (Bell and Boardman, 1992; Butzer, 1982; Dotterweich, 2013; Goudie, 2013; Hillel, 1991; Kennett and Beach, 2013; Kirch, 2005; Marston, 2015; McNeill and Winiwarter, 2006; Montgomery, 2007; Sandor and Homburg, 2015). A historic example is the Dust Bowl in the US southern Great Plains, where extensive clearing and cultivation of prairie combined with a periodic drought led to severe wind erosion, land abandonment, and depopulation (Egan, 2006; Montgomery, 2007). Climate change, now linked with human causes, presents serious challenges to societies today. The failures and successes of ancient societies in their land use and response to a wide range of climate and environmental change can provide long-term insights valuable in addressing current problems such as water supply and agricultural production under difficult and variable climate conditions.

CONCLUDING STATEMENTS Many scientists now consider us to be living in the Anthropocene, an age when human influences on the planetary system have become so pronounced that humans are the single most defining geologic force of the era (Steffen et al., 2015; Zalasiewicz et al., 2015). This influence has made many soils humannatural bodies rather than just natural bodies (Yaalon and Yaron, 1966; Richter and Yaalon, 2012; Brevik et al., 2016b). Another impact of humans on the natural system has been our influence on changes to global climate (IPCC, 2014). Soils are both changed by climate and, in turn, influence climate (Brevik, 2012). Therefore, as we seek to understand what living in a world of changing climate may mean for modern civilizations, understanding how soils will respond and how that affects humans is also a pressing need (Hungate et al., 2003; Zhang et al., 2004; Grace et al., 2006; Brevik, 2009; Ping et al., 2015; Keesstra et al., 2016). One approach to undertaking this task is to model what the future

References 21 will look like given the changes that we expect to see. However, human civilizations have lived under changing climates in the past. Through interdisciplinary and transdisciplinary studies that link archaeologists, climate scientists, economists, social scientists, soil scientists, and others who can contribute to a holistic understanding of the impact of climate change on past human societies and the natural resources, such as soils, that they depended on, we have the opportunity to observe those impacts. By paying attention to and learning from our past, we can improve the models that seek to predict our future and better plan to manage potential impacts.

ACKNOWLEDGMENTS E.C. Brevik was partially supported by the National Science Foundation EPSCoR program under Grant Number IIA-1355466 during this project.

REFERENCES Adams, K.R., 2015. The archaeology and agronomy of ancient maize (Zea Mays L.). In: Ingram, S.E., Hunt, R.C. (Eds.), Arid Lands Agriculture: Understanding the Past for the Future. University of Arizona Press, Tucson, pp. 15e53. Amundson, R., Berhe, A.A., Hopmans, J.W., Olson, C., Sztein, A.E., Sparks, D.L., 2015. Soil and human security in the 21st century. Science 348, 1261071. ´ .E., Arneborg, J., Heinemeier, J., Lynnerup, N., Nielsen, H.L., Rud, N., Sveinbjo¨rnsdo´ttir, A 1999. Change of diet of the Greenland Vikings determined from stable carbon isotope analysis and 14C dating of their bones. Radiocarbon 41 (2), 157e168. Aubet, M.E., 2008. Political and economic implications of the new Phoenician chronologies. In: Sagona, C. (Ed.), Beyond the Homeland: Markers in Phoenician Chronology. Peeters Publishers, Leuven, pp. 179e191. Barlow, L.K., Sadler, J.P., Ogilvie, A.E.J., Buckland, P.C., Amorosi, T., Ingimundarson, J.H., Skidmore, P., Dugmore, A.J., McGovern, T.H., 1997. Interdisciplinary investigations of the end of the Norse Western Settlement in Greenland. The Holocene 7 (4), 489e499. Barrera-Bassols, N., Toledo, V.M., 2005. Ethnoecology of the Yucatec Maya: symbolism, knowledge and management of natural resources. Journal of Latin American Geography 4 (1), 9e41. Bell, M., Boardman, J. (Eds.), 1992. Past and Present Soil Erosion: Archaeological and Geographical Perspectives. Oxbow Books, Oxford. Benson, L.V., Berry, M.S., 2009. Climate change and cultural response in the prehistoric American Southwest. KIVA 75, 89e117. Benson, L.V., Pauketat, T.R., Cook, E.R., 2009. Cahokia’s boom and bust in the context of climate change. American Antiquity 74, 467e483. Berglund, J., 1986. The decline of the Norse settlements in Greenland. Arctic Anthropology 23 (1&2), 109e135. Binford, M.W., Kolata, A.L., Brenner, M., Janusek, J.W., Seddon, M.T., Abbott, M., Curtis, J.H., 1997. Climate variation and the rise and fall of an Andean civilization. Quaternary Research 47, 235e248. Boucher de Perthes, J., 1847. Antiquite´s celtiques et ante´diluviennes. Me´moire sur l’industrie primitive et les arts a` leur origine. Cambridge University Press, Cambridge. Branch, N.P., Kemp, R.A., Silva, B., Meddens, F.M., Williams, A., Kendall, A., Pomacanchari, C.V., 2007. Testing the sustainability and sensitivity to climatic change of terrace agricultural systems in the Peruvian Andes: a pilot study. Journal of Archaeological Science 34, 1e9.

22


Brevik, E.C., 2009. Soil, food security, and human health. In: Verheye, W. (Ed.), Soils, Plant Growth and Crop Production. Encyclopedia of Life Support Systems (EOLSS). EOLSS Publishers, Oxford. http://www.eolss.net. Brevik, E.C., 2012. Soils and climate change: gas fluxes and soil processes. Soil Horizons 53 (4), 12e23. Brevik, E.C., Fenton, T.E., 2012. Long-term effects of compaction on soil properties along the Mormon Trail, south-central Iowa, USA. Soil Horizons 53 (5), 37e42. Brevik, E.C., Hartemink, A.E., 2010. Early soil knowledge and the birth and development of soil science. Catena 83, 23e33. Brevik, E.C., Fenton, T.E., Homburg, J.A., 2016a. Historical highlights in american soil science e prehistory to the 1970s. Catena 146, 111e127. https://doi.org/10.1016/ j.catena.2015.10.003. Brevik, E.C., Richter, D.B., Verrecchia, E.P., Ryan, J., Poch, R.M., Crouvi, O., Sauer, D., Waroszewski, J., Solleiro-Rebolledo, E., Monger, C., Ottner, F., Targulian, V., 2016b. The influence of Dan H. Yaalon: his impact on people. Catena 146, 147e154. https:// doi.org/10.1016/j.catena.2015.10.012. Buckland, P.C., Amorosi, T., Barlow, L.K., Dugmore, A.J., Mayewski, P.A., Mcgovern, T.H., Ogilvie, A.E.J., Sadler, J.P., Skidmore, P., 1996. Bioarchaeological and climatological evidence for the fate of Norse farmers in medieval Greenland. Antiquity 70, 88e96. Bull, W.B., 1991. Geomorphic Responses to Climatic Change. Oxford University Press, New York. Buringh, P., 1960. Soils and Soil Conditions in Iraq. Ministry of Agriculture, Republic of Iraq, Baghdad. Butzer, K.W., 1982. Archaeology as Human Ecology: Method and Theory for a Contextual Approach. Cambridge University Press, Cambridge. Cao, Z.H., Ding, J.L., Hu, Z.Y., Knicker, H., Ko¨gel-Knabner, I., Yang, L.Z., Yin, R., Lin, X.G., Dong, Y.H., 2006. Ancient paddy soils from the Neolithic age in China’s Yangtze River Delta. Naturwissenschaften 93, 232e236. Coffey, G.N., 1911. The development of soil survey work in the United States with a brief reference to foreign countries. Proceedings of the American Society of Agronomy 3, 115e129. Critchley, W.R.S., Reij, C., Willcocks, T.J., 1994. Indigenous soil and water conservation: a review of the state of knowledge and prospects for building on traditions. Land Degradation & Development 5, 293e314. Dean, J.S., 1988a. Dendrochronology and paleoenvironmental reconstruction on the Colorado Plateau. In: Gummerman, G.J. (Ed.), The Anasazi in a Changing Environment. Cambridge University Press, Cambridge, pp. 119e167. Dean, J.S., 1988b. A model of Anasazi behavioral adaption. In: Gummerman, G.J. (Ed.), The Anasazi in a Changing Environment. Cambridge University Press, Cambridge, pp. 25e44. de Menocal, P.B., 2001. Cultural responses to climate change during the late Holocene. Science 292, 667e673. Denevan, W.M., 2001. Cultivated Landscapes of Native Amazonia and the Andes. Oxford University Press, New York. Diamond, J., 2005. Collapse: How Societies Choose to Fail or Succeed. Viking, New York. Doolittle, W.E., 2000. Cultivated Landscapes of Native North America. Oxford University Press, New York. Dotterweich, M., 2013. The history of human-induced soil erosion: geomorphic legacies, early descriptions and research, and the development of soil conservation-a global synopsis. Geomorphology 201, 1e34. Doyel, D.E., Dean, J.S. (Eds.), 2006. Environmental Change and Human Adaptation in the Ancient American Southwest. University of Utah Press, Salt Lake City. Drake, B.L., Wills, W.H., Erhardt, E.B., 2012. The 5.1 ka aridization event, expansion of pinõn-juniper woodlands, and the introduction of maize (Zea mays) in the American Southwest. The Holocene 22 (12), 1353e1360.

References

23

Dunning, N.P., Beach, T., Farrell, P., Luzzadder-Beach, S., 1998. Prehispanic agrosystems and adaptive regions in the Maya lowlands. Culture and Agriculture 20 (2/3), 87e101. Dunning, N.P., Beach, T.P., Luzzadder-Beach, S., 2012. Kax and kol: collapse and resilience in lowland Maya civilization. Proceedings of the National Academy of Sciences 109, 3652e3657. Egan, T., 2006. The Worst Hard Time: The Untold Story of Those Who Survived the Great American Dust Bowl. Houghton Mifflin Co., Boston. Entwistle, J.A., 2000. The geoarchaeological significance and spatial variability of a range of physical and chemical soil properties from a former habitation site, Isle of Skye. Journal of Archaeological Science 27, 287e303. Fagan, B.M., 2008. The Great Warming: Climate Change and Rise and Fall of Civilizations. Bloomsbury Press, New York. Fernańdez-Romero, M.L., Lozano-Garcıá, B., Parras-Alcańtara, L., 2014. Topography and land use change effects on the soil organic carbon stock of forest soils in Mediterranean natural areas. Agriculture, Ecosystems & Environment 195, 1e9. Fitzpatrick, R.W., 2004. Changes in Soil and Water Characteristics of Natural, Drained and Re-flooded Soils in the Mesopotamian Marshlands: Implications for Land Management Planning. CSIRO Land and Water Client Report. http://www.clw.csiro.au/publications/ consultancy/2004/Mesopotamian-marshlands-soil.pdf. Fitzpatrick, R.W., 2013. Demands on soil classification and soil survey strategies: specialpurpose soil classification systems for local practical use. In: Developments in Soil Classification, Land Use Planning and Policy Implications. Springer, Netherlands, pp. 51e83. Giamrnellaro, A.S., 1999. Phoenician agriculture and its produce. In: Flandrin, J.L., Montanari, M. (Eds.), Food: A Culinary History from Antiquity to the Present. Columbia University Press, New York, pp. 55e65. Glaser, B., Birk, J.J., 2014. State of the scientific knowledge on properties and genesis of Anthropogenic Dark Earths in Central Amazonia (terra preta de Ińdio). Geochimica et Cosmochimica Acta 82, 39e51. Gong, Z., Zhang, X., Chen, J., Zhang, G., 2003. Origin and development of soil science in ancient China. Geoderma 115, 3e13. Goudie, A.S., 2013. The Human Impact on the Natural Environment: Past, Present, and Future, seventh ed. Wiley-Blackwell, Chichester. Grace, P.R., Colunga-Garcia, M., Gage, S.H., Robertson, G.P., Safir, G.R., 2006. The potential impact of agricultural management and climate change on soil organic carbon of the north central region of the United States. Ecosystems 9, 816e827. Grayson, D., 1983. The Establishment of Human Antiquity. Academic Press, New York. Hambach, U., Rolf, C., Schnepp, E., 2008. Magnetic dating of Quaternary sediments, volcanites and archaeological materials: an overview. Quaternary Science Journal 57 (1e2), 25e51. Hamidian, A., Ghorbani, M., Abdolshahnejad, M., Abdolshahnejad, A., 2015. Qanat, traditional eco-technology for irrigation and water management. Agriculture and Agricultural Science Procedia 4, 119e125. Hardman, J.R., Heuberger, D.C., Collins, M.E., 1998. Characterization of soils at the Coralie archaeological site Grand Turk, British West Indies. Soil Survey Horizons 39 (1), 1e15. Hart, J.P., Brumbach, H.J., Lusteck, R., 2007. Extending the phytolith evidence for early maize (Zea mays ssp. mays) and squash (Cucurbita sp.) in central New York. American Antiquity 72 (3), 563e583. Heckenberger, M.J., Kuikuro, A., Kuikuro, U.T., Russell, C.J., Schmidt, M., Fausto, C., Franchetto, B., 2003. Amazonia 1492: pristine forest or cultural parkland. Science 301, 1710e1713. Hellin, J., Haigh, M.J., 2002. Better land husbandry in Honduras: towards the new paradigm in conserving soil, water and productivity. Land Degradation & Development 13, 233e250. Heyerdahl, T., 1961. An introduction to easter island. In: Heyerdahl, T., Ferdon, E.N. (Eds.), Reports of the Norwegian Archaeological Expedition to Easter Island and the East Pacific. Archaeology of Easter Island, vol. 1. Forum Publishing House, Stockholm, pp. 21e92.

24


Hillel, D., 1991. Out of the Earth: Civilization and the Life of the Soil. University of California Press, Berkeley. Hoggarth, J.A., Breitenbach, S.F.M., Culleton, B.J., Ebert, C.E., Masson, M.A., Kennett, D.J., 2016. The political collapse of Chicheń Itza´ in climatic and cultural context. Global and Planetary Change 1368, 25e42. Holliday, V.T., 1992. Soil formation, time, and archaeology. In: Holliday, V.T. (Ed.), Soils in Archaeology: Landscape Evolution and Human Occupation. Smithsonian Institution Press, Washington, DC, pp. 101e117. Holliday, V.T., 2004. Soils in Archaeological Research. Oxford University Press, New York. Homburg, J.A., 1988. Archaeological investigations at the LSU campus mounds. Louisiana Archaeology 15, 31e204. Homburg, J.A., 2005. Archaeology in relation to soils. In: Hillel, D., Rosenweig, C., Powlson, D., Scow, K., Singer, M., Sparks, D. (Eds.), Encyclopedia of Soils and the Environment. Elsevier, Oxford, pp. 95e102. Homburg, J.A., Sandor, J.A., 2011. Anthropogenic effects on soil quality of ancient agricultural systems of the American Southwest. Catena 85, 144e154. Homburg, J.A., Douglas, D.L., Brevik, E.C., Tepley, C., Orme, A., 2014a. Paleoenvironmental background. In: Homburg, J.A., Douglass, J.G., Reddy, S.N. (Eds.), People in a Changing Land: The Archaeology and History of the Ballona in Los Angeles, California. Statistical Research, Inc., Tucson, pp. 73e84. Homburg, J.A., Douglas, D.L., Brevik, E.C., Tepley, C., Orme, A., 2014b. Paleoenvironmental reconstruction of the Ballona Lagoon. In: Homburg, J.A., Douglass, J.G., Reddy, S.N. (Eds.), People in a Changing Land: The Archaeology and History of the Ballona in Los Angeles, California. Statistical Research, Inc., Tucson, pp. 85e109. Homburg, J.A., Sandor, J.A., Norton, J.B., 2005. Anthropogenic influences on Zuni agricultural soils. Geoarchaeology 20 (7), 661e693. Huber, E.K., 2004. Early maize at the old Corn site (LA 137258). In: Huber, E.K., Van West, C.R. (Eds.), Fence Lake Archaeological Project: Archaeological Data Recovery in the New Mexico Transportation Corridor and First Five-Year Permit Area, Fence Lake Coal Mine Project, Catron County, New Mexico, Synthetic Studies and Summary. Statistical Technical Series, vol. 4. Statistical Research, Inc., Tucson, pp. 36.1e36.14. Huckleberry, G., 2015. Defining the environmental context of indigenous agriculture in the Southwest: what we don’t know about Middle to Late Holocene climate change and floodplain dynamics. In: Ingram, S.E., Hunt, R.C. (Eds.), Arid Lands Agriculture: Understanding the Past for the Future. University of Arizona Press, Tucson, pp. 89e130. Hungate, B.A., Dukes, J.S., Shaw, M.R., Luo, Y., Field, C.B., 2003. Nitrogen and climate change. Science 302, 1512e1513. https://doi.org/10.1126/science.1091390. Hunt, T.L., 2007. Rethinking Easter Island’s ecological catastrophe. Journal of Archaeological Science 34 (3), 485e502. Hunt, T.L., Lipo, C.P., 2006. Late colonization of easter island. Science 311, 1603e1606. Ibań˜ez, J.J., Pe´rez-Go´mez, R., Oyonarte, C., Brevik, E.C., 2015. Are there arid land soilscapes in southwestern Europe? Land Degradation & Development 26, 853e862. Ingram, S.E., 2015. Human vulnerability to dry periods. In: Ingram, S.E., Hunt, R.C. (Eds.), Arid Lands Agriculture: Understanding the Past for the Future. University of Arizona Press, Tucson, pp. 131e160. IPCC, 2014. Climate change 2014: synthesis report. In: Core Writing Team, Pachauri, R.K., Meyer, L.A. (Eds.), Contribution of Working Groups I, II and III to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. IPCC, Geneva. Jacob, J.S., 1995. Archaeological pedology in the Maya lowlands. In: Pedological Perspectives in Archeological Research, SSSA Special Publication 44. Soil Science Society of America, Madison, pp. 51e80. Keesstra, S.D., Bouma, J., Wallinga, J., Tittonell, P., Smith, P., Cerda`, A., Montanarella, L., Quinton, J.N., Pachepsky, Y., van der Putten, W.H., Bardgett, R.D., Moolenaar, S., Mol, G., Jansen, B., Fresco, L.O., 2016. The significance of soils and soil science towards realization of the United Nations Sustainable Development Goals. SOIL 2, 111e128. https://doi.org/10.5194/soil-2-111-2016.

References 25 Kennett, D.J., Beach, T.P., 2013. Archeological and environmental lessons for the Anthropocene from the Classic Maya collapse. Anthropocene 4, 88e100. Khaledian, Y., Kiani, F., Ebrahimi, S., Brevik, E.C., Aitkenhead-Peterson, J., 2017. Assessment and monitoring of soil degradation during land use change using multivariate analysis. Land Degradation and Development 28, 128e141. https://doi.org/10.1002/ ldr.2541. Kirch, P.V., 2005. Archaeology and global change: the Holocene record. Annual Review of Environment and Resources 30, 409e440. King, F.H., 1911. Farmers of Forty Centuries. Privately published by Mrs. F.H. King. http:// rodare.cl/fda/unidad-i/FarmersofFortyCenturies.pdf. Krupenikov, I.A., 1992. History of Soil Science from Its Inception to the Present. Oxonian Press, New Delhi. Lal, R., 2007. Evolution of the plow over 10,000 years and the rationale for no-till farming. Soil and Tillage Research 93, 1e12. Lal, R., 2009. Laws of sustainable soil management. Agronomy for Sustainable Development 29, 7e9. Landa, E.R., Brevik, E.C., 2015. Soil science and its interface with the history of geology community. Earth Sciences History 34 (2), 296e309. Larson, G., Piperno, D.R., Allaby, R.G., Purugganan, M.D., Andersson, L., Arroyo-Kalin, M., Barton, L., Vigueira, C.C., Denham, T., Dobney, K., Doust, A.N., Gepts, P., Gilbert, M.T.P., Gremillion, K.J., Lucas, L., Lukens, L., Marshall, F.B., Olsen, K.M., Pires, J.C., Richerson, P.J., de Casas, R.R., Sanjur, O.I., Thomas, M.J., Fuller, D.Q., 2014. Current perspectives and the future of domestication studies. Proceedings of the National Academy of Sciences of the United States of America 111 (17), 6139e6146. Li, C., Cao, S., 1990. History of soil science in China. Transactions of the 14th International Congress of Soil Science 5, 197e202. Lombardo, U., Denier, S., Veit, H., 2015. Soil properties and pre-Columbian settlement patterns in the Monumental Mounds Region of the Llanos de Moxos, Bolivian Amazon. SOIL 1, 65e81. https://doi.org/10.5194/soil-1-65-2015. Macphail, R.I., Cruise, G.M., Allen, M.J., Linderholm, J., Reynolds, P., 2004. Archaeological soil and pollen analysis of experimental floor deposits; with special reference to Butser Ancient Farm, Hampshire, UK. Journal of Archaeological Science 31, 175e191. Mahaffy, J.P., 1889. The work of Mago on agriculture. Hermathena 7 (15), 29e35. Mairura, F.S., Mugendi, D.N., Mwanje, J.I., Ramisch, J.J., Mbugua, P.K., Chianu, J.N., 2008. Scientific evaluation of smallholder land use knowledge in central Kenya. Land Degradation & Development 19, 77e90. Mann, D., Edwards, J., Chase, J., Beck, W., Reanier, R., Mass, M., Finney, B., Loret, J., 2008. Drought, vegetation change, and human history on Rapa Nui (Isla de Pascua, Easter Island). Quaternary Research 69, 16e28. Marston, J.M., 2015. Modeling resilience and sustainability in ancient agricultural systems. Journal of Ethnobiology 35, 585e605. May, D.W., Hill, M.G., Holven, A.C., Loebel, T.J., Rapson, D.J., Semken Jr., H.A., Theler, J.L., 2008. Geoarchaeology of the Clary Ranch Paleoindian sites, Western Nebraska. In: Raynolds, R.G. (Ed.), Roaming the Rocky Mountains and Environs: Geological Field Trips, Geological Society of America Field Guide 10. Geological Society of America, Boulder, pp. 265e293. Mayer, J.H., Surovell, T.A., Waguespack, N.M., Kornfeld, M., Reider, R.G., Frison, G.C., 2005. Paleoindian environmental change and landscape response in Barger Gulch, Middle Park, Colorado. Geoarchaeology 20 (6), 599e625. McMichael, A.J., 2003. Global climate change and health: an old story writ large. In: Climate Change and Human Health: Risks and Responses. World Health Organization, Geneva, pp. 1e17. McNeill, J.R., Winiwarter, V. (Eds.), 2006. Soils and Societies: Perspectives from Environmental History. White Horse Press, Isle of Harris.

26


Meeks, S.C., Anderson, D.G., 2013. Drought, subsistence stress, and population dynamics. In: Wingard, J.D., Hayes, S.E. (Eds.), Soils, Climate, and Society. University Press of Colorado, Boulder, pp. 61e83. Miller, B.A., Schaetzl, R.J., 2014. The historical role of base maps in soil geography. Geoderma 230e231, 329e339. Minnis, P. (Ed.), 2014. New Lives for Ancient and Extinct Crops. University of Arizona Press, Tucson. Monger, H.C., 1995. Pedology in arid lands archaeological research: an example from Southern New Mexico-Western Texas. In: Pedological Perspectives in Archeological Research, SSSA Special Publication 44. Soil Science Society of America, Madison, pp. 35e50. Montgomery, D.R., 2007. Dirt: The Erosion of Civilizations. University of California Press, Berkeley. Moseley, M.E., 2001. The Incas and Their Ancestors: The Archaeology of Peru. Thames and Hudson, London. Nabhan, G.P., 2013. Growing Food in a Hotter, Drier Land: Lessons from Desert Farmers on Adapting to Climate Uncertainty. Chelsea Green Publishing, White River Junction. Nelson, M.C., Hegmon, M., Kintigh, K.W., Kinzig, A.P., Nelson, B.A., Anderies, J.M., Abbott, D.A., Spielmann, K.A., Ingram, S.E., Peeples, M.A., Kulow, S., Strawhacker, C.A., Meegan, C., 2012. Long-term vulnerability and resilience: three examples from archaeological study in the southwestern United States and northern Mexico. In: Cooper, J., Sheets, P. (Eds.), Surviving Sudden Environmental Change: Answers from Archaeology. University Press of Colorado, Boulder, pp. 197e220. Neves, E.G., Petersen, J.B., Bartone, R.N., Da Silva, C.A., 2003. Historical and socio-cultural origins of Amazonian Dark Earths. In: Lehman, J., Kern, D.C., Glaser, B., Woods, W.I. (Eds.), Amazonian Dark Earths: Origin, Properties, Management. Kluwer Academic Publishers, Dodrecht, pp. 29e50. Parnell, J.J., Terry, R.E., Sheets, P., 2002. Soil chemical analysis of ancient activities in Cereń, El Salvador: a case study of a rapidly abandoned site. Latin American Antiquity 13 (3), 331e342. Ping, C.L., Jastrow, J.D., Jorgenson, M.T., Michaelson, G.J., Shur, Y.L., 2015. Permafrost soils and carbon cycling. SOIL 1, 147e171. https://doi.org/10.5194/soil-1-147-2015. Powell, M.A., 1985. Salt, seed, and yields in Sumerian agriculture: a critique of the theory of progressive salinization. Zeitschrift fur Assyrologie 75, 7e38. Pringle, H., 1997. Death of Norse in Greenland. Science 275, 924e926. Rengasamy, P., 2006. World salinization with emphasis on Australia. Journal of Experimental Botany 57 (5), 1017e1023. Richter, D., 2007. Humanity’s transformation of Earth’s soil: pedology’s new frontier. Soil Science 172 (12), 957e967. Richter, D., Yaalon, D.H., 2012. “The changing model of soil” revisited. Soil Science Society of America Journal 76 (3), 766e778. Rost, S., 2015. Watercourse Management and Political Centralization in Third-millennium B.C. Southern Mesopotamia: A Case Study of the Umma Province of the Ur III Period (2112e2004 B.C.) (unpublished Ph.D. dissertation). State University of New York at Stony Brook, Stony Brook. Rutherford, G.K., 1995. Soils of some Norse settlements in southwestern Greenland. Arctic 48 (4), 324e328. Sandor, J.A., 2006. Ancient agricultural terraces and soils. In: Warkentin, B. (Ed.), Footprints in the Soil: People and Ideas in Soil History. Elsevier, Amsterdam, pp. 505e534. Sandor, J.A., Burras, C.L., Thompson, M., 2005. Human impacts on soil formation. In: Hillel, D. (Ed.), Encyclopedia of Soils in the Environment, vol. 1. Elsevier, Oxford, pp. 520e532. Sandor, J.A., Eash, N.S., 1991. Significance of ancient agricultural soils for long-term agronomic studies and sustainable agriculture research. Agronomy Journal 83, 29e37. Sandor, J.A., Eash, N.S., 1995. Ancient agricultural soils in the Andes of southern Peru. Soil Science Society of America Journal 59, 170e179.

References 27 Sandor, J.A., Homburg, J.A., 2011. Soil and landscape responses to American indian agriculture in the American Southwest. In: Nelson, M.C., Strawhacker, C.A. (Eds.), Movement, Connectivity, and Landscape Change. University Press of Colorado, Boulder, pp. 141e159. Sandor, J.A., Homburg, J.A., 2015. Agricultural soils of the prehistoric Southwest: known unknowns. In: Ingram, S.E., Hunt, R.C. (Eds.), Arid Lands Agriculture: Understanding the Past for the Future. University of Arizona Press, Tucson, pp. 54e88. Sandor, J.A., Gersper, P.L., Hawley, J.W., 1986. Soils at prehistoric agricultural terracing sites in New Mexico, pts. 1e3. Site placement, soil morphology, and classification. Organic matter and bulk density changes. Phosphorus, selected micronutrients, and pH. Soil Science Society of America Journal 50, 173e180. Sandor, J.A., WinklerPrins, A.M.G.A., Barrera-Bassols, N., Zinck, J.A., 2006. The heritage of soil knowledge among the world’s cultures. In: Warkentin, B. (Ed.), Footprints in the Soil: People and Ideas in Soil History. Elsevier, Amsterdam, pp. 43e84. Sarris, A., Galaty, M.L., Yerkes, R.W., Parkinson, W.A., Gyucha, A., Billingsley, D.M., Tate, R., 2004. Geophysical prospection and soil chemistry at the early Copper age settlement of Ve´sztT-Bikeri, southeastern Hungary. Journal of Archaeological Science 31, 927e939. Scarborough, V.L., 2003. The Flow of Power: Ancient Water Systems and Landscapes. SAR Press, Santa Fe. Schiffer, M.B., 1987. Formation Process of the Archaeological Record. University of New Mexico Press, Albuquerque. Schmidt, A., 2007. Archaeology, magnetic methods. In: Gubbins, D., Herrero-Bervera, E. (Eds.), Encyclopedia of Earth Sciences Series, Encyclopedia of Geomagnetism and Paleomagnetism. Springer, New York, pp. 23e31. Schwindt, D.M., Bocinsky, R.K., Ortman, S.G., Glowacki, D.M., Varien, M.D., Kohler, T.A., 2016. The social consequences of climate change in the central Mesa Verde region. American Antiquity 81, 74e96. Simpson, I.A., Barrett, J.H., Milek, K.B., 2005. Interpreting the Viking age to medieval period transition in Norse Orkney through cultural soil and sediment analyses. Geoarchaeology 20 (2), 357e379. Smekalova, T., Voss, O., Abrahamsen, N., 1993. Magnetic investigation of iron-smelting centres at Snorup, Denmark. Archaeologia Polona 31, 83e103. Sparks, D.L., 2006. Historical aspects of soil chemistry. In: Warkentin, B.P. (Ed.), Footprints in the Soil: People and Ideas in Soil History. Elsevier, Amsterdam, pp. 307e337. Steffen, W., Broadgate, W., Deutsch, L., Gaffney, O., Ludwig, C., 2015. The trajectory of the Anthropocene: the great acceleration. The Anthropocene Review 2 (1), 81e98. Sweetwood, R.V., Terry, R.E., Beach, T., Dahlin, B.H., Hixson, D., 2009. The Maya footprint: soil resources of Chunchucmil, Yucatań, Mexico. Soil Science Society of America Journal 73, 1209e1220. Tisdale, S.L., Nelson, W.L., Beaton, J.D., Havlin, J.L., 1993. Soil Fertility and Fertilizers, fifth ed. MacMillan Publishing Company, New York. Troeh, F.R., Hobbs, J.A., Donahue, R.L., 2004. Soil and Water Conservation for Productivity and Environmental Protection, fourth ed. Prentice Hall, Upper Saddle River. Vasey, D.E., 2002. An Ecological History of Agriculture 10,000 BC to AD 10,000. Purdue University Press, West Lafayette. Warren, D.M., 1994. Indigenous agricultural knowledge, technology, and social change. In: McIsaac, G., Edwards, W.R. (Eds.), Sustainable Agriculture in the American Midwest: Lessons from the Past, Prospects for the Future. University of Illinois Press, Urbana, pp. 35e53. Waters, M.R., 2006. Prehistoric human response to landscape change in the American Southwest. In: Doyel, D.E., Dean, J.S. (Eds.), Environmental Change and Human Adaptation in the Ancient American Southwest. University of Utah Press, Salt Lake City, pp. 26e45. Weiss, H., Bradley, R.S., 2001. What drives societal collapse? Science 291, 609e610.

28


Weiss, H., Courty, M.-A., Wetterstrom, W., Guichard, F., Senior, L., Meadow, R., Curnow, A., 1993. The genesis and collapse of third millennium north Mesopotamian civilization. Science 261, 995e1004. Wilson, C., Simpson, I.A., Currie, E.J., 2002. Soil management in pre-Hispanic raised field systems: micromorphological evidence from Hacienda Zuleta, Ecuador. Geoarchaeology 17 (3), 261e283. Winiwarter, V., 2006. Soil scientists in ancient Rome. In: Warkentin, B.P. (Ed.), Footprints in the Soil: People and Ideas in Soil History. Elsevier, Amsterdam, pp. 3e16. Yaalon, D.H., Arnold, R.W., 2000. Attitudes toward soils and their societal relevance: then and now. Soil Science 165 (1), 5e12. Yaalon, D.H., Yaron, B., 1966. Framework for man-made soil changes-an outline of metapedogenesis. Soil Science 102 (4), 272e277. Zalasiewicz, J., Waters, C.N., Williams, M., Barnosky, A.D., Cearreta, A., Crutzen, P., Ellis, E., Ellis, M.A., Fairchild, I.J., Grinevald, J., Haff, P.K., Hajdas, I., Leinfelder, R., McNeill, J., Odada, E.O., Poirier, C., Richter, D., Steffen, W., Summerhayes, C., Syvitski, J.P.M., Vidas, D., Wagreich, M., Wing, S.L., Wolfe, A.P., An, Z., Oreskes, N., 2015. When did the Anthropocene begin? A mid-twentieth century boundary level is stratigraphically optimal. Quaternary International 383, 196e203. Zhang, X.C., Nearing, M.A., Garbrecht, J.D., Steiner, J.L., 2004. Downscaling monthly forecasts to simulate impacts of climate change on soil erosion and wheat production. Soil Science Society of America Journal 68, 1376e1385.

Chapter | Two

SoilePlanteAtmosphere Interactions: Ecological and Biogeographical Considerations for Climate-Change Research Lucas C.R. Silva*, 1, Hans Lambersx

*Environmental Studies Program, Department of Geography, Institute of Ecology & Evolution, University of Oregon, Eugene, Oregon, United States; xSchool of Plant Biology, Faculty of Natural and Agricultural Sciences, The University of Western Australia, Crawley, Australia 1


INTRODUCTION Long before the development of the scientific method, our ancestors’ survival depended on their ability to identify the distribution patterns of terrestrial organisms. Through observation, they learned that certain characteristics of climate and landscape position coincide with geographical boundaries within which plants and animals occur. Through experimentation, they expanded those natural boundaries and intensified the production of useful species which fueled the development of ancient civilizations into modern societies. It was not until recently, however, that the global consequences of human activity became apparent. Despite numerous technological advances in modern history, demand for natural resources has increased and so has the impact of our species on land, air, and water across the globe. It is now clear that the human influence on land cover and fossil fuel emissions have caused changes in atmospheric composition and climate that reverberate through natural and managed ecosystems worldwide. To understand whether and how these pressures can be absorbed by the biosphere, it will serve us well to consider processes that have controlled ecosystem responses to global environmental change since the emergence of life on land. 29 Climate Change Impacts on Soil Processes and Ecosystem Properties. https://doi.org/10.1016/B978-0-444-63865-6.00002-8 Copyright © 2018 Elsevier B.V. All rights reserved.

30

SoilePlanteAtmosphere Interactions

For millions of years, life has thrived at the critical zone between Earth’s slow-changing crust and its volatile atmosphere. The boundaries within which terrestrial organisms occur are delineated by climate and by the availability of limiting resources that cycle through the soil-plant-atmosphere (SPA) interface. In this chapter, we explore how symbiotic associations that emerged to overcome resource limitation in ancient environments still shape the composition and distribution of terrestrial ecosystems today. We discuss climate-induced ecosystem transformations that occurred before and after human activity became a dominant planetary force and evaluate the connection between ecological and biogeographical regime shifts and climatic stability. We conclude by identifying gaps in knowledge and research directions that will improve climate-change prediction and impact-mitigation efforts in natural and managed ecosystems. Many uncertainties surround the evolutionary trajectories of terrestrial organisms and forces of diversification and assembly that organize species into the complex ecosystem configurations we see today. There are, however, at least three undisputable points relevant to that story. First, plants and all other organisms that depend upon them (including humans) rely on resources derived from two distinct sources, the Earth’s crust and its atmosphere. The primary source of mineral nutrients needed for plant growth is the soil, which is formed from bedrock through biological activity, and the primary source of energy that fuels biological activity is the organic material generated via photosynthetic assimilation of carbon dioxide (CO2) from the air. Second, ecosystem assembly involves nonrandom processes. Terrestrial organisms respond to their environment by adapting and co-operating (or competing) to overcome resource limitation. As a result, species associations vary across gradients of resource availability in which levels of biodiversity tend to overlap with the diversity of resource-acquisition strategies. Third, terrestrial organisms not only respond to their environment but also influence environmental conditions. The influence of organisms on the environment spans multiple scales, from the regulation of local hydrological and biogeochemical cycles to the stabilization of the global climate. In the following sections, each of these points will be discussed in detail. For now, the central premise to keep in mind is that the ability of terrestrial ecosystems to absorb the impact of environmental change depends on intricate processes that emerged in ancient environments and still sustain life at the SPA interface today. For at least half a century, the SPA interface has been studied as an integrated system (Philip, 1966). Readers of this book are probably familiar with the term SPA “continuum,” which has been traditionally used to describe water movement from the soil through plants and into the air in production systems (Nova´k, 2012). Recent developments have brought a new dimension to research in this field, as insights from natural ecosystems redefined the unidirectional assumption of a “continuum” to include movement of mass and energy in the opposite direction; i.e. from the air through plants and into the soil (e.g., Oliveira et al., 2005; Limm et al., 2009; Eller et al., 2013; Earles et al., 2016). Interest in the

Terrestrial Life as a Stabilizing Climatic Force

31

bidirectional transfer of mass and energy at the SPA interface has been stimulated by the need to understand and scale interdependent fluxes of water, carbon, and nutrients from individual organisms to the globe (e.g., Bonan et al., 2014; Schlesinger and Jasechko, 2014; Evaristo et al., 2015). Promising areas of investigation arising from this pursuit include the study of hydrological and biogeochemical cycles as a means to anticipate ecosystem transformations that are important for human well-being (e.g., Berhe and Kleber, 2013; Dieleman et al., 2015; Grau et al., 2017; Pecl et al., 2017). Here, we review wellestablished concepts and discuss new developments that provide a path for advancing research into SPA interactions as an integral component of terrestrial ecosystem dynamics. Background for this exploration is provided by the central theme of this book, which is concerned with the impact of climatic change on terrestrial ecosystems. Within this broad theme we focus on specific mechanisms that are rooted in evolutionary history and, at the same time, that will be useful for guiding new conservation, restoration, and management efforts. Unlike previous anthologies of SPA-related literature, which tend to emphasize physical processes, we focus on biogeochemical processes that can aid the prediction of ecosystem transformations and their influence on climatic stability. Through the lenses of ecology and biogeography, we identify knowledge gaps that can be addressed by the integrated analysis of SPA interactions across a broad range of scales and disciplines.

TERRESTRIAL LIFE AS A STABILIZING CLIMATIC FORCE The history of complex biological systems began on Earth at the onset of the Phanerozoic eon, a period characterized by the presence of “visible” (phanero´sd4anεrό2) “life” (z oḗdzuή), which spans over 540 million years, from the appearance of the first multicellular eukaryotic organisms to modern human-dominated environments. Through its journey toward complexity, life confronted major challenges, some of which can be identified in fossil and molecular records of mass extinctions (Rohde and Muller, 2005; Whiteside and Grice, 2016). Of the five largest known extinction events, one is attributed to the impact of meteorites (Schulte et al., 2010) and the others, which occurred during periods of particularly strong volcanic activity (Bond and Wignall, 2014), have been attributed to changes in climate that followed the release of greenhouse gases from igneous provinces (Whiteside and Grice, 2016). Evidently, life endured despite recurrent extinction events and, over time, diversified in ways that stabilized Earth’s climate. During much of the Phanerozoic, global temperature oscillations were decoupled from atmospheric CO2 levels. With the exception of the Carboniferous, a carbon-bearing period that lasted from w360 to 300 million years before present (YBP), the estimated concentration of atmospheric CO2 ranged from 2 to 15 times above the current level; yet, large temperature oscillations including cold and warm cycles occurred (Fig. 2.1). As evolution progressed,

32


FIGURE 2.1 Temperatures, CO2, and biodiversity curves Adapted from Silva & Anand 2013. Estimated emergence of plant-microbe associations (Brundrett, 2009; Field et al., 2015); root evolution (Kenrick and Strullu-Derrien 2014); angiosperm radiation (Friis et al., 2011, 2016); ice-free Arctic oceans (Vandermark et al., 2009); tropical plants in Greenland (Boyd, 1992).

the amplitude and timing of fluctuations in the temperature and CO2 fluctuations converged. High-resolution records of atmospheric CO2 based on analyses of soil carbonate and fossil leaves show that even the most intense and prolonged cold interval in Earth’s history (the late Paleozoic ice age; w300 million YBP) encompassed multiple glacialeinterglacial cycles driven by feedbacks between vegetation productivity and atmospheric CO2 concentration (Montan˜ez et al., 2016). Synchronous fluctuations are apparent both at coarse scales (e.g., 10-million-year intervals) and fine resolutions (e.g., million-year intervals), with CO2 trajectories varying in concert with global temperature anomalies and biogeographical rearrangements, particularly toward the late Phanerozoic (Royer, 2006). The persistent warming of the planetary system that followed the Paleozoic ice age is attributed to volcanic CO2 emissions, which led to global temperatures w6 C higher than today’s average (Fig. 2.1) and coincided with the Permian extinction (w250 million YBP; Bond and Wignall, 2014). This warm interval was followed by a cold period during the Triassic and a long warm period that extended into the Cretaceous (w140e60 million YBP). This latter warm interval was characterized by the presence of marine life in ice-free Arctic oceans (Vandermark et al., 2009) and tropical plants in areas that are presently under glaciers (e.g., Greenland; Boyd, 1992). Notably, this period also marked a steep increase in terrestrial biodiversity that coincided with the beginning of a persistent decline in atmospheric CO2 levels, in synchrony with the earliest known global angiosperm radiation w100 million YBP (Fig. 2.1). The strength of the connection between levels of biodiversity and terrestrial carbon sinks remains a matter of debate even in contemporary systems (e.g., Ladd et al., 2012; Brienen et al., 2015; Sheil et al., 2016; Sullivan et al., 2017). However, we find broad support for the hypothesis that the

Terrestrial Life as a Stabilizing Climatic Force

33

diversification and expansion of terrestrial organisms functioned as a stabilizing climatic force throughout history. For example, arbuscular mycorrhizal associations that emerged as rudimentary root systems w400 million YBP are thought to have accelerated bedrock weathering and soil formation at continental scales, removing vast amounts of carbon from the atmosphere (Brundrett, 2009; Kenrick and Strullu-Derrien, 2014). Carbon sequestration was magnified by the diversification of vascular plants and expansion of terrestrial ecosystems borne by adaptive innovations that included root ericoid and ectomycorrhizal associations, capable of recycling complex organic molecules (Read and Perez-Moreno, 2003) and carboxylate-exuding cluster roots, capable of enhancing mineral nutrient acquisition highly-weathered nutrient-poor soil horizons (Lambers et al., 2015). Interestingly, the steepest known decline in atmospheric CO2 concentration occurred at the onset of the Devonian (i.e. beginning of the Carboniferous) when roots became a global biogeochemical force (Raven and Edwards, 2001), driving accumulating terrestrial carbon at increasingly greater depths, and leaching of carbonates to the ocean (Royer et al., 2004; Brundrett, 2009; Kenrick and Strullu-Derrien, 2014). In the recent geological past, fluctuations in temperature and CO2 concentration have been strongly correlated, as evidenced in temperature and CO2 records from ice cores that span the past eight glacialeinterglacial cycles (w800,000 YBPdEpica, 2004; Lu¨thi et al., 2008). The primary driver of glaciale interglacial climate oscillations is the variation in Earth’s orbit eccentricity and obliquity of its axis (i.e. Milankovitch cycles), which result in periodic changes in species ranges and reorganization of terrestrial biomes (Svenning et al., 2015). However, changes in biogeographical boundaries associated with glacialeinterglacial cycles also function as a biospheric “thermostat”d i.e. increasing terrestrial carbon sequestration during warm periods and releasing additional CO2 to the atmosphere during cold periods (Archer, 2010). As a result, the amplitude of climatic oscillations in the recent geological past is orders of magnitude smaller than those reported for the early- to mid-Phanerozoic (Prentice et al., 2011; Blois et al., 2013). Given that every ecosystem that has persisted through time must have acquired some degree of resilience to fluctuations in climate and atmospheric composition, modern biomes can be expected to continue functioning as a stabilizing climatic force. It is important to note, however, that preindustrial fluctuations in climate allowed time for species to adapt to new environments and shift their optimal range, whereas the velocity of climate warming caused by land-use change and burning of fossil fuels since the industrial revolution has no historical analogue (IPCC, 2013). As we enter the Anthropocene, a geological epoch marked by humanity’s influence on Earth (Schimel et al., 2013), climatic change coupled with land-use pressures can impede species migration and gene flow, threatening the persistence of many ecosystems or disrupting their climatic buffering effect. Indeed, the exceptionally rapid loss of biodiversity observed over the past century alone suggests that a sixth mass extinction is under way (Ceballos et al., 2015). In all likelihood, the forces of natural selection will continue to influence how organisms and ecosystems respond to human-

34


BOX 2.1 Recent Past Recent ecosystem rearrangements driven by climate occurred at the beginning of the Holocene, about 12,000 years ago. Since then, warm postglacial climates have supported vegetation forms that are similar to those that exist today. In northern regions, early Holocene ecosystem range shifts caused by warming can be inferred from fossil pollen records, which reveal a widespread postglacial expansion of forests over grasslands (Williams et al., 2004). The distribution of tropical and subtropical ecosystems was mostly stable during that period, but since w5000 YBP evergreen (wet) and deciduous (dry) forests have been expanding and retreating, respectively (Ledru, 2002; Silva et al., 2008, 2010; Pessenda et al., 2009). The response of tropical systems implies a delayed increase in water availability that allowed forests to expand over savannas and grasslands despite increasing human occupation and fire disturbance during that period (Behling and Pillar, 2007; Taylor et al., 2010; Pinter et al., 2011; Carson et al., 2014). In South America, multiple lines of evidence including the reconstructed timing of early human colonization from charcoal particles, carbon isotope signatures of soil organic matter, and fossil pollen suggest that the expansion of forests resulted from an intensification of monsoonal precipitation (Ledru, 2002; Marchant et al., 2002; Silva and Anand, 2011; 2013). This interpretation is supported by independent climatic records from the Andes (e.g., Bird et al., 2011b), western Amazon basin (Apae´stegui et al., 2014), and central Brazil (Wortham et al., 2017). There appears to exist a latitude-dependent pattern in the timing of forest expansion that suggest variable ecological sensitivities at different distances from the monsoon core region (Silva, 2014), which is consistent with expected increases in moisture from high to low latitudes as a result of melting ice caps during the early- to mid-Holocene warming (Lamy and Kaiser, 2009). This period was also followed by a decline (w30 ppm) in atmospheric CO2 concentration (Fluckiger et al., 2002), presumably caused by an overall increase in terrestrial productivity. Increased bedrock weathering in other regions of the Southern (Dixon et al., 2016) and Northern Hemispheres (Brovkin et al., 2012; Fritz and Anderson, 2013) likely contributed to this decline in CO2 concentration. Indeed, the effect of changes in water regime and vegetation distribution on parent material weathering show the potential for up to fourfold increases in weathering from ancient environments to recently established ecosystems (Taylor et al., 2012). These estimates indicate a strong temperature control on the kinetics of mineral dissolution, but root exudates and rootemicrobe associations are known to enhance weathering rates beyond those that can be attributed solely to climate (Moulton et al., 2000; Berner, 2004; Taylor et al., 2009).

induced global environmental change. Thus, the study of evolutionary legacies and symbiotic associations that control the distribution and function of terrestrial ecosystems will continue to be important for predicting, and perhaps mitigating, climatic change and its ecological reverberations in the future (Box 2.1).

CONTEMPORARY SYSTEMS Parallels of Soil and Ecosystem Development Interactions between past and present SPA-mediated adaptive processes create a dynamic balance between opposing forces of soil and ecosystem formation or loss. For example, in warm and humid regions, soils undergo pervasive weathering losses and chemical depletion of essential mineral elements (Vitousek, 2004). In contrast, atmospheric inputs of mineral elements often exceed weathering losses in cold and arid regions due to low solute transport (Chadwick et al., 1999; Bristow et al., 2010). At the same time, climate influences species

Contemporary Systems 35 distribution (Whittaker, 1975; Bartlein et al., 2011) and the activity of soil organisms (Bond-Lamberty and Thomson, 2010), thereby, controlling the net input of organic material into the soil (Bonan, 2008; Beer et al., 2010) as well as the recycling of essential nutrients at the SPA interface (Field et al., 2007; Potter et al., 2008). In other words, Soil profiles are active mixing zones of living and dead organic materials and decaying minerals, where chemical reactions provide ligands for carbon stabilization, a process that diminishes the loss of mineral nutrients through leaching (Chadwick and Chorover, 2001; Chadwick et al., 2003; Rasmussen et al., 2008). In turn, soil properties strongly influence the physiological performance of dominant plant species as well as the spatiotemporal aggregation of diverse communities that ultimately control how ecosystems respond to climatic variability. The net result of these SPA

FIGURE 2.2 (A) Global map of soil orders (USDA-NRCS). (B) Vascular plant diversity and distribution of endemic species - endemism richness (ER) range equivalents per 10,000 km2 - which generally overlap with the distribution of soil orders, with notable exceptions for Mediterranean regions. Reprinted with permission from Kier, G., Kreft, H., Lee, T.M., et al., 2009. A global assessment of endemism and species richness across island and mainland regions. Proceedings of the National Academy of Sciences of the United States of America 106, 9322e9327.

36


associations is apparent in the overlapping distribution of natural vegetation and soil types across the globe (Fig. 2.2A and B). The parallels of edaphic and ecological trajectories have long been expressed as a function of a few common factors of formation, where: soil/ecosystem ¼ f (climate, topography, parent material, organisms, time). In this classic model, all soil- or ecosystem-forming factors are treated as independent variables, a corollary being that each factor can be studied separately when the others are held constant (Dokuchaev, 1883; Jenny, 1941; Walker and Syers, 1976; Amundson and Jenny, 1997). This model has been useful for understanding how soils and ecosystems change in response to environmental conditions and will continue to be useful as it is revised to incorporate human impact and its interactions with other forming factors (Amundson and Sposito, 2013). Like other organisms, humans appear in the role of dependent as well as independent factor of soil and ecosystem formation. Human dependency is illustrated by examples of past civilizations that failed to prevent degradation of the soils on which they were founded (Diamond, 2011) and by modern cases of declining productivity due to soil erosion across the globe (FAO, 2012). On the other hand, human activities function collectively as an independent forming factor that has altered soils and ecosystems, causing biotic intactness to decline beyond 10%dthe proposed “safe” threshold for sustainabilitydacross 65% of the Earth’s land surface (Newbold et al., 2016). Beyond human impacts, the single most important predictor of species diversity across modern terrestrial biomes is kinetics; i.e. the energy dependence of ecological and evolutionary transformations (Brown, 2014). Not unlike soil development, global biodiversity patterns and rates of speciation across modern biomes are strongly and positively correlated with historical climatic stability and its corresponding time-integrated energy inputs (Hopper, 2009; Fine, 2015). At regional scales, biotic interactions maintain levels of biodiversity, such that positive associations emerge between soil development and functional species diversity along resource gradients that capture the influence of other soil formation factors. As a result, biodiversity hotspots are commonly found in regions of high geodiversity (i.e. variety of geological features; Fig. 2.2B). Globally, 35 biodiversity hotspots with more than 3000 plant species per hectare are found in warm and humid regions where highly weathered soils commonly occur, but there are also Mediterranean and seasonally dry tropical regions where species diversity is extremely high (Myers et al., 2000; Mittermeier et al., 2011; Williams et al., 2011). Notable examples include native vegetation forms of South Africa, central Brazil, south-western Australia, and Central Chile (Myers et al., 2000). These are among the regions that cover a very small percentage of the Earth’s land, and yet harbor thousands of plant species (Barthlott et al., 2007; van der Ent and Lambers, 2016), many of which are restricted to severely impoverished soils (Oliveira et al., 2015; Silveira et al., 2016). In regions where ancient soils prevail, edaphic diversity is subtle and resource gradients generated by plant communities themselves are thought to play a major role in maintaining levels of biodiversity. For example, plant species adapted to

Contemporary Systems 37 either tropical forest or savanna environments have markedly distinct functional traits, associated with specialized resource-use and growth strategies, which have been shown to reinforce nutrient gradients through differential uptake and litter deposition that favor the persistence of multiple stable states under homogeneous climatic and lithologic conditions (Rossatto et al., 2009; Hoffmann et al., 2012; Silva et al., 2013a; Staal et al., 2016).

Bottom-Up Regulation of Symbiotic Associations Different edaphic environments show different amplitudes of restriction for resource acquisition, such that the diversity of species and of resource acquisition strategies tend to increase with soil age (Fig. 2.3). As soon as fresh bedrock is exposed, microbial communities take possession of it and initiate pedogenesis (Brown and Jumpponen, 2015). At early stages of soil development, nitrogenfixing species dominate, but as soil development progresses, nitrogen content increases and primary productivity shifts from being nitrogen limited to being increasingly limited by mineral nutrients, such as phosphorus and calcium (Vitousek et al., 2010; Vitousek and Chadwick, 2013). At intermediate stages of soil development, marked shifts occur from states in which arbuscular mycorrhizas make up most of the microbial biomass binding soil particles (Smith et al., 2008; Kallenbach et al., 2016) to states dominated by ectomycorrhizas capable of accessing organic nutrient pools that are not available to arbuscular fungi

FIGURE 2.3 Resource-acquisition strategies and plant diversity trajectory across soil-age gradients. Adapted from Lambers, H., Raven J.A., Shaver, G.R., Smith, S.E. 2008. Plant nutrientacquisition strategies change with soil age. Trends Ecology & Evolution 23, 95e103. and Zemunik, G., Turner, B.L., Lambers, H., Laliberte´, E., 2016. Increasing plant species diversity and extreme species turnover accompany declining soil fertility along a long-term chronosequence in a biodiversity hotspot. Journal of Ecology 104, 792e805.

38


(Lambers et al., 2008). Nonmycorrhizal plants predominantly occur on old soils where specialized carboxylate-releasing root clusters can mine otherwise unavailable mineral nutrients (Zemunik et al., 2015, 2016). Notable examples include species adapted to the oldest and most impoverished soils of southwestern Australia, which in addition to root adaptations exhibit a remarkable mineralnutrient-resorption proficiency (Hayes et al. 2014) and show toxicity symptoms when exposed to slightly elevated levels of soil fertility (Lambers et al., 2010). Similarly, in ancient landscapes of central Brazil, where all major nutrients are lacking and levels of aluminum, manganese, and iron are toxic for most crops, many native species accumulate toxic elements in leaves, whereas others do not survive in the absence of exchangeable aluminum, even though no specific role for that element has been documented in plant metabolism (Haridasan, 2008; Guilherme Pereira et al., 2018). The effect of symbiotic associations on soil formation and carbon pools is often independent of and, in some cases, larger than the effects of climate or mineralogy. Approximately 80% of all contemporary plant species form mycorrhizal symbioses, with arbuscular mycorrhizal associations being the most common (Brundrett, 2009). Globally, soil in ecosystems that are dominated by ectomycorrhizal and ericoid mycorrhizal fungi contains 70% more carbon per unit nitrogen than soil in ecosystems that are dominated by arbuscular mycorrhizal fungi (Averill et al., 2014). It follows that proportionally more carbon can be sequestered in ectomycorrhizal and ericoid mycorrhizal systems under nitrogen limitation. Moreover, symbiotic associations exert a stronger effect on bedrock weathering than climate does. This is true both for ancestral arbuscular-mycorrhizal associations and for more recent lineages of ectomycorrhizal fungi and cluster roots (Taylor et al., 2009). For the contemporary climate and CO2 levels, mycorrhizal symbioses increase continental weathering by a factor of 1.3, reaching 1.9 for sites dominated by evergreen needle-leaved trees, and 2.8 for sites dominated by deciduous broad-leaved trees (Taylor et al., 2012). Weathering by carboxylate-releasing cluster roots as a means to mobilize phosphate (Hinsinger, 2001) or the release of iron-chelating phytosiderophores can lead to even faster weathering rates (Robin et al., 2008). Furthermore, the effects of warming and elevated CO2 on ecosystems vary depending on symbiotic strategies (Fernandez et al., 2016; Terrer et al., 2016) and on species traits that optimize productivity under nutrient limitation, particularly in highly-weathered soils (e.g., Asner et al., 2016a; Bahar et al., 2016). In such soils, biomolecular phosphorus and carboxyl groups facilitate microbial adhesion to iron oxides, which can lead to levels of carbon accumulation that go well beyond expected lithological thresholds of organic matter stabilization (e.g., Parikh et al., 2014; Silva et al., 2015b). Symbiotic associations also control interplant resource transfer through roote microbe networks, which have been shown to mediate complex adaptive responses of plant communities to environmental change (Gorzelak et al., 2015). Such responses include long-distance transport of carbon and nutrients via advective mass flow through common mycorrhizal networks as well as active transport through mycelium growth (Simard et al., 2012). Accumulating empirical

Simplifying Complexity at the SoilePlanteAtmosphere Interface

39

evidence support the notion that the productivity and regenerative capacity of ecosystems under environmental stress depend on this type of mutualistic resource-use strategy. Notable examples include the flow of carbon from old trees to regenerating seedlings through mycorrhizal connections, which follow photosynthetic gradients that favor transfer to seedlings on disturbed soils (Teste et al., 2010). Similarly, water transfer from water-replete to drought-stressed plants through mycorrhizal networks have been shown to buffer the impact of drought stress beyond individual organisms in temperate forests (Bingham and Simard, 2011, 2012). We have only recently begun to understand how these symbiotic associations, which are costly in terms of carbon demand, scale from organisms to ecosystems. On the one hand, plants that are “hubs” for mycorrhizal networks can “even out” stand-level resource availability (Simard, 2009), creating favorable conditions for sustained productivity via transfer of photosynthates between plants and microbial communities whose carbon can persist in the soil for millennia (e.g., Trumbore, 2009). On the other hand, depending on soil age and lithology mycorrhizal networks can amplify competition by, for example, preferentially allocating mineral nutrients to large host plants (Weremijewicz et al., 2016) and accelerating litter decomposition as well as soil carbon and nitrogen loss (Qiao et al., 2014; Winsome et al., 2017). Although many uncertainties still exist with respect to the resilience-building power of root-microbe interactions in terrestrial ecosystems, it is clear that any realistic approach to predicting ecological responses to changing climates must consider SPA-mediated mutualistic or competitive resource acquisition strategies.

SIMPLIFYING COMPLEXITY AT THE SOILePLANTeATMOSPHERE INTERFACE In an attempt to summarize the various processes discussed above, a simplified model of climatically relevant SPA feedbacks is presented in Fig. 2.4. Soil organic carbon is a central term in this conceptual model. Given that there would not be soil without plants, in this model vegetation drives soil development by providing carbon inputs through litter deposition, root growth, and rhizodeposition of organic compounds. Thus, shifts in vegetation composition, distribution, or productivity exert a direct effect on soil development. Microorganisms also play an important role in soil formation. Root exudates and rootemicrobe associations control the rates at which bedrock weathering progresses and vertical soil development occurs. Physical and chemical transformations of the original bedrock, in turn, regulate the potential for plant and microbial growth as well as soil carbon saturation thresholds. Closing the loop, the concentration of atmospheric CO2 is positively associated with temperature and tends to stimulate primary productivity. Increases in productivity can abate CO2 emissions, functioning as a negative force on temperature; however, significant complexities surround the effect of climate and CO2 on plant growth and soil development, which also depend on variable water regimes, which remain difficult to predict (dashed arrows; Fig. 2.4).

40


d

Atmospheric CO2 a

Vegetation composition & productivity

h

Temperature b

c i e Root exudates & symbiotic associations

Soil organic carbon

g

j f

Bedrock weathering

FIGURE 2.4 Conceptual model of major climatically-relevant soil-plant-atmosphere interaction. Solid arrows represent direct effects (e.g. rising CO2 levels lead to higher temperatures). Grey squares indicate inverse effects (e.g. increasing temperatures lead to decreasing soil carbon stocks). A positive (i.e. destabilizing) feedback loop links atmospheric CO2 concentration, temperature, and soil carbon content (a-b-c). A negative (i.e. stabilizing) feedback loop links atmospheric CO2, vegetation productivity, and rootemicrobe associations (d-e-f), which affect and are influenced by soil carbon content (j). Dashed arrows represent uncertain feedbacks involving changes in water regime and their effects on vegetation and soil formation (g-h-i). Scheme adapted from Amundson, R., Berhe, A.A., Hopmans, J.W., et al., 2015. Soil science. Soil and human security in the 21st century. Science (New York, N.Y.) 348, 1261071 and Taylor, L.L., Banwart, S.A., Valdes, P.J., Leake, J.R., Beerling, D.J., 2012. Evaluating the effects of terrestrial ecosystems, climate and carbon dioxide on weathering over geological time: a global-scale process-based approach. Philosophical Transactions of the Royal Society of London. Series B, Biological Sciences 367, 565e582.

The stabilizing effect of SPA interactions on climate is primarily determined by the balance between carbon inputs, plant and microbial metabolic losses, and biogeochemical mechanisms of bedrock weathering and organic matter stabilization. Currently, the soil carbon pool is estimated to be over three times larger than the atmospheric pool (Que´re´ & et al., 2016). Thus, although the magnitude and stability of terrestrial carbon sinks remain the subject of controversy (e.g., Lal, 2008; Berhe and Kleber, 2013; Smith et al., 2016; Ko¨rner, 2017; Silva, 2017; Wang et al., 2017), any ecological or biogeographical shifts that have the potential to affect soil carbon stocks will also have a significant effect on climate. It is important to note that changes in species richness do not necessarily correspond to changes in the distribution of functional traits that govern how ecosystems respond to environmental change (Enquist et al., 2015). However, there is great potential for improving predictions of ecosystem responses to climate and CO2 concentration based on patterns of species diversity and productivity. Overall, the relationship between functional trait diversity and climate is modest, but some striking patterns emerge when resource-use strategies are considered (e.g., leaf economic spectrum; Wright et al., 2004). Changes in species functional traits and phenology affect ecosystem carbon and water cycles (Polgar and Primack, 2011; Bonan et al., 2011) as well as

Simplifying Complexity at the SoilePlanteAtmosphere Interface

41

the ability of roots to acquire nutrients through mass flow and diffusion (Elmore et al., 2016; Maxwell et al., 2018a). Accordingly, the relative abundance of lowefficiency species or traits that correspond to fast water or nutrient uptake can be used to translate individual physiological performance into collective community- or ecosystem-level mass and energy fluxes (Reich, 2012, 2014). In regions where climate warming is associated with increasing soil water and nutrient availability plant growth and ecosystem productivity are expected to increase is response to elevated CO2. This has been the case in many alpine ecotones where permafrost thaw and enhanced snowmelt coupled with warming-induced shifts in phenology have caused recent increases in tree growth and forest cover (e.g.,; Salzer et al., 2009; Ko¨rner, 2012; Silva et al., 2016). Other examples include temperate regions across western North America, where the net effect of warming and elevated CO2 has resulted in enhanced tree growth and evapotranspiration, outpacing decreases in stomatal conductance from elevated CO2 (Mankin et al., 2017). Similarly, in many warm tropical and subtropical regions C3 woody species have recently expanded into areas previously dominated by C4 grasses (Franco et al., 2014; Stevens et al., 2016), which is consistent with anticipated positive effects of CO2 enrichment on the less efficient C3 relative to C4 metabolic pathways (Ainsworth and Long, 2005), as the latter evolved to optimize photosynthesis in low-CO2 environments (Edwards and Still, 2008). Globally, warming- and CO2-induced land “greening” has led to an overall increase in land-to-air water loss through transpiration (Zhang et al., 2015). Although elevated CO2 generally increases the amount of CO2 fixed per unit of water transpired (Keenan et al., 2013; van der Sleen et al., 2014), it also results in decreased nutrient flow from the soil matrix to plants. For this reason, the benefits of CO2-induced increases in water-use efficiency are often outweighed by the negative effect of water and nutrient limitation, which have caused widespread growth decline and mortality of dominant tree species in many drought-prone regions (e.g., Williams et al., 2012; Go´mez-Guerrero et al., 2013;Asner et al., 2016b). A fundamental tradeoff between efficiency and productivity intersects multiple scales of biological organization and can be leveraged to simplify complexities and improve ecological predictions at the SPA interface. At the molecular level, the kinetics of release and sorption of different nutrients in soil can be used to predict the rate of enzymatic CO2 reduction and subsequent transport of soluble sugars from photosynthesizing sites through plants and into the rhizosphere (e.g., Shi et al., 2015; Sperling et al., 2017). At the organism level, knowledge of the amount and chemical form of nutrients in the soil solution (e.g., ammonium or nitrate) can be used to estimate changes in net productivity as well as above-to-belowground biomass allocation of trees under elevated CO2 (Silva et al., 2015a). At the community level, assessments of climate-driven changes in species composition on colocated infertile and fertile soils have shown high levels of ecological stability in infertile soils in temperate grasslands (Harrison et al., 2015) and in tropical forestesavanna ecotones (Paiva et al., 2015). This occurs because resource limitation favors the dominance of conservative traits and resource-efficient growth strategies at the

42


expense of productivity (Hopper, 2009; Zemunik et al., 2015; Silveira et al., 2016). At the ecosystem scale, the sensitivity of forests and grasslands to elevated CO2 is limited by soil water and nutrient availability, such that resource limitation increases following an initial CO2-induced stimulation of ecosystem productivity (Finzi et al., 2006; Izquierdo et al., 2013; Reich et al., 2014). Similarly, the response of managed forests and agricultural systems to changes in climate and CO2 levels is subjected to tradeoffs involving productivity and soil water and nutrient uptake. For example, an analysis of irrigated and nonirrigated lands show that nutrient management must be taken into account to predict yield responses to climate and CO2 concentration (Shaw et al., 2014). Furthermore, a meta-analysis of CO2-enrichment experiments using staple crops of the modern human diet show that the benefits of CO2 stimulation of productivity and water-use efficiency are counteracted by declines in nutritional value (e.g., elevated carbon-to-protein ratios; Myers et al., 2014). Thus, although feedbacks between water regime and plant physiology can change the direction and magnitude of relationships between carbon and nutrient cycles (e.g., Creed et al., 2014; Saha and Setegn, 2015), the conceptual framework proposed here can be used to guide future assessments of climate and CO2 concentration effects on natural and managed lands.

GAPS IN KNOWLEDGE Significant progress has been made to represent the range and diversity of resource -acquisition strategies that can be predicted on the basis of soil properties and dominant plant species (Lambers et al., 2018). However, we still lack a quantitative understanding of how different symbiotic associationsdsome more costly than others with respect to carbon demand (Raven et al., 2018)d scale from local to regional scales. Mathematical models developed to predict the spatial overlap (i.e. co-occurrence) between species subject to positive interactions (e.g., mutualism and commensalism) or negative interactions (e.g., competition and amensalism) generate discernible patterns that challenge the widely held view that climate alone is sufficient to characterize species distributions (Arau´jo and Rozenfeld, 2013). Nevertheless, the biogeochemical drivers and consequences of such interactions remain poorly defined. Foremost among issues to be addressed is the quantitative characterization of mechanisms controlling unexpected effects of climate and atmospheric composition on the functioning of ecosystems. It is well known that complex and nonlinear SPA interactions affect planetary energetics through biogeochemical and biogeophysical feedbacks (Bonan, 2008; Anderson et al., 2011), but it is unclear whether and how seemingly weak connections between fast, locally controlled biological processes (e.g., leaf gas exchange and rootemicrobe associations) and slow, spatially broad geophysical and geochemical processes (e.g., bedrock weathering and soil development) can be used to predict changes in ecosystem distribution and function. Major gaps in knowledge include uncertain changes in hydrological cycles (Fig. 2.4) and biogeochemical mechanisms of plantemicrobial interactions, which could potentially be addressed by recent developments integrated SPA research (Box 2.2).

Gaps in Knowledge

43

BOX 2.2 Promising Areas for Future Research BiophysicaldBeyond carbon sequestration SPA interactions contribute to at least two other major terrestrial climate forcings: radiative balance (dependent on land cover and surface albedo) and evaporative cooling (dependent on evapotranspiration) (Law et al., 2002; Bonan, 2008; Williams et al., 2012b). Not unlike carbon fluxes, these climate forcings are governed by fast, locally controlled processes (e.g., leaf transpiration) and slow, spatially broad processes (e.g., species range shifts). A variety of global ecological datasets can now be used in process-based models capable of predicting shifts in species composition and function (Evans et al., 2016) and progress has been made in recent years to quantify and scale corresponding changes in terrestrial carbon and water fluxes (e.g., through eddy covariance sensors; Baldocchi, 2005; Monson and Baldocchi, 2014 and remote sensing techniques Asner et al., 2016a,b). However, we still lack historical baselines for hydrological changes. For example, the characterization of an “unequivocal warming of the climate system [that is] unprecedented over decades to millennia” contrasts with a tentative account of hydrological shifts, which are described as “likely affected by anthropogenic influence” (IPCC, 2013). This difference in language reflects different levels of confidence in paleotemperature records, for which we have reliable baselines, in contrast with poorly understood water regimes. Isotopic assessments of modern biomes indicate that over 80% of the global terrestrial water loss to the atmosphere happens through transpiration (Jasechko et al., 2013). This estimate varies widely with location and data treatment (Coenders-Gerrits et al., 2014), but even in the most conservative scenario shows that plant-mediated water fluxes exert a major influence on the global water cycle (Schlesinger and Jasechko, 2014). A recent warming-induced intensification of evapotranspiration can be inferred from runoff data (Alkama et al., 2011; Destouni et al., 2012), but plant cover and soil properties also influence water losses, generating negative (stabilizing) or positive (reinforcing) feedbacks that are difficult to model (Baldocchi, 2014). In managed agricultural and forest ecosystems, stable isotope analysis of ancient seeds and tree rings can help elucidate historical changes in water regime and their associated effects on carbon and nutrient cycles (e.g., Maxwell et al., 2014; Riehl et al., 2014). From individual plants to entire ecosystems, plantderived cellulose and hemicellulose-degradation products found in soil profiles can be used as records of the balance between precipitation and evapotranspiration (Tuthorn et al., 2014; Zech et al., 2014). Recent work on bulk lipid extracts (Silva et al., 2015c) and more refined compound-specific analysis (Sachse et al., 2012; Maxwell et al., 2018b) promise to improve reconstructions of water regime shifts and their ecological and biogeographical consequences. BiogeochemicaldWe have just begun to understand how costly symbiotic associations can improve plant growth and increase ecosystem resilience to environmental change. Large amounts of limiting soil resources are transferred via rootemicrobe networks (Gorzelak et al., 2015; Song et al., 2015) and the connection between biodiversity and energy-limited symbiotic associations that govern nutrient acquisition is clear (Fig. 2.3). However, the broad impact of ericoid and ectomycorrhizas, which are capable of accessing nutrients from recalcitrant organic complexes, and of cluster roots or functionally similar root specializations, which are capable of using carboxylates to release unavailable soil minerals, remains poorly defined (Lambers et al., 2015). These associations have been described as a “blind spot” in the cost-benefit analysis of soileplant relations (Bol et al., 2016) and could help explain patterns of species diversity and ecosystem stability in nutrient-poor environments. In addition, other nutrient sources have the potential to influence the cost of symbiotic associations and the fate of carbon on land. In some nitrogen-limited ecosystems carbon storage has been shown to be associated with the distribution of nitrogen-rich bedrock (Morford et al., 2011, 2016) and variation in the chemical form of soil nitrogen can affect how plants respond to rising CO2 levels (Bloom et al., 2010). Interactions between soil and atmospheric sources of nitrogen (e.g., ammonia; Silva et al., 2015a) can also cause significant differences in plant productivity. However, these and other biological processes remain absent in ecosystem and Earth system Continued

44


BOX 2.2 Promising Areas for Future ResearchdContinued models. For example, asymbiotic (i.e. free-living) microorganisms are capable of alleviating nutrient limitation in impoverished soils. Typically, the contribution of free-living nitrogen-fixing species is less than that of symbiotic associations, but depending on the developmental stage of an ecosystem, asymbiotic nutrient inputs can be larger and more evenly distributed than symbiotic inputs (Sullivan et al., 2014). As in other biological processes, the concentration of mineral elements (e.g., phosphorus, iron, and molybdenum) control the activity of free-living organisms (Reed et al., 2011). Thus, as ecosystem development progresses, the immobilization of nutrients into organic material constrains asymbiotic nitrogen fixation (Pe´rez et al., 2004). To date, most climate models do not account for biogeochemical SPA interactions. For this reason, climate predictions tend to overestimate the rates of carbon sequestration on land and underestimate the pace of global warming (He et al., 2016). Inconsistencies between models and observations have at least three additional causes. First, warming-induced stress can limit the positive effect of rising CO2 levels on primary productivity. Second, the effect of plantemicrobe association on organic matter stabilization and parent material weathering is generally not included in climate predictions. Third, the production and emission of greenhouse gases, other than CO2, in natural ecosystems and the fate of fertilizers used in agricultural systems have not been sufficiently described. There is a growing need for explicitly representing these and other aspects of SPA-mediated processes in Earth-system models (Wieder et al., 2015). Empirical case studies that point to promising research directions include: (1) relationships between microbial functional genes and enzyme activities that show how carbon degradation can be predicted by gene abundance for improved models of carbon sequestration (Trivedi et al., 2016); (2) the effect of atmospheric nutrient deposition on biological nitrogen fixation (Yu et al., 2015) and on phosphorus and other mineral inputs (Gross et al., 2016) that affect biodiversity (Bobbink et al., 2010); and (3) the influence of vegetation cover on methanotroph and methanogen microbial communities and seasonal patterns of CH4 emissions (Bridgham et al., 2012; Rodrigues et al., 2013; Meyer et al., 2017; Morris et al., 2017). Finally, future studies stand to gain valuable information by considering abiotic drivers of global biogeochemical cycles, such as sunlight-induced reactions involving components of land, air, and water, which include the production of greenhouse gases and the release of essential elements needed for primary production (Doane, 2017).

Although grounded in systems thinking, the SPA component of Earth sciences is rarely studied as a complex system. Most of what we know about ecosystems’ response to climate comes from correlative models of temperature and vegetation distribution, which do not account for biologically controlled feedbacks involving resource consumption or environmental modulation of tradeoffs that govern species assembly (Higgins, 2017). For this reason, the response of the terrestrial biosphere to increasing atmospheric CO2 concentration is a major source of uncertainty in models that project future climate change scenarios (e.g., Arneth et al., 2010; Friedlingstein et al., 2014). Stateof-the-art models are still unable to reproduce observed changes in climate and vegetation (Harrison et al., 2016), resulting in major discrepancies between

Conservation and Management Opportunities

45

modelled and observed alterations of terrestrial carbon and water cycles (Smith et al., 2016). Experimental determination of climate-induced changes in plant communities and biogeochemical cycles have improved our mechanistic understanding of ecosystem dynamics (e.g., Yospin et al., 2015; Pfeifer-Meister et al., 2016), but comparisons of experimental data and dynamic vegetation models demonstrate a need for new theory capable of integrating biological processes at the SPA interface (Medlyn et al., 2015). For example, plant transpiration is a major factor controlling water fluxes from the land to the atmosphere (e.g., Jasechko et al., 2013), but using this well-known concept in a predictive manner necessitates knowledge of how soil properties affect the physiological performance of plant species. Soil properties not only influence the intrinsic wateruse efficiency and mass flow of nutrients from the soil matrix to the root of individual trees (Maxwell et al. 2018a) but also exert a bottom-up control on symbiotic associations (Fig. 2.3) and on the distribution of vegetation through landscapes of high lithologic diversity (e.g., Hahm et al., 2014). Soil properties are, however, slow to respond compared to climate-induced shifts in tree physiology or the composition of plant communities. Therefore, a hierarchical approach is needed to understand how dynamic and inertial ecosystem properties interact. Specifically, a combination of ecological, physiological, and biogeochemical principles is necessary to scale what is known about local level processes to the global environment (Silva, 2015).

CONSERVATION AND MANAGEMENT OPPORTUNITIES The global soil carbon reservoir today is estimated to be w1500 Pg (1 Pg C ¼ 1012 kg of C), which is greater than the pools of carbon found in the atmosphere and all plant life combined, but depleted relative to prehistoric levels (Houghton, 2007; Oelkers and Cole, 2008; Que´re´ & et al., 2016). The estimated loss of soil organic carbon due to human impacts since the onset of plant and soil domestication during the Neolithic, ranges from 40 to 100 Pg C (Joosten, 2015). If temperature projections hold, global soil carbon stocks are expected to fall w55 Pg C by 2050, which is between 12% and 17% of the total anthropogenic CO2 emissions projected for that period (Crowther et al., 2016). The economic value of soil carbon inputs for reaching 2050 storage targets ranges from US$350 billion to several trillion dollars each year (Loftus et al., 2015). Between 15% and 37% of terrestrial species living today in carbon-rich areas that cover w20% of the Earth’s surface could be “committed to extinction” by 2050 (Thomas et al., 2004), and yet the yearly expenditure on global biodiversity conservation is approximately US$21 billion, which is less than a third of the estimated amount needed to achieve biodiversity conservation targets (Sheil et al., 2016). Thus, better integration of climate and land-use change policies is needed to mitigate carbon and biodiversity losses and to improve the return on investment of funds used to address local management and conservation priorities.

46


Multiple benefits can be achieved through improved land conservation and management that focus on SPA interactions that are capable of increasing carbon sequestration in species-rich areas. For example, tropical forests cover only w10% of the global land area (Fig. 2.5A), but harbor the most plant and animal species on Earth (Fig. 2.2) and are responsible for w35% of the global primary productivity and for >25% of the terrestrial carbon stock (Fig. 2.5B and C). In addition, tropical forests provide an additional cooling effect due to evapotranspiration rates that are much higher than those of savannas, grasslands, or agricultural systems (Zeri et al., 2014; Nobre et al., 2016; Jerszurki et al., 2017), despite having lower albedo than those open vegetation physiognomies. Temperate forests, on the other hand, are responsible for w12% of the global primary production, but hold only w10% of the terrestrial carbon stock and are often considered a carbon source due to continued deforestation (Bonan, 2008). Nevertheless, carbon stocks of temperate forests are much larger than

(A)

(B)

(D)

Aboveground

Belowground

(C)

FIGURE 2.5 Global area (A), above- and belowground (to a depth of 3 m) carbon stock (B), and net primary productivity (C) for major forests and nonforest biomes, including agricultural systems and deserts (D). Reprinted with permission from Silva, L.C.R., Anand, M., 2013. Historical links and new frontiers in the study of forest-atmosphere interactions. Community Ecology 14, 208e218.

Final Considerations 47 those of temperate grasslands or croplands (Fig. 2.5D), and thus conservation and expansion of temperate forests would attenuate CO2 emissions. The boreal region stores large amounts of carbon in permafrost soils and wetlands. Boreal forest expansion has been closely associated with rising global temperatures, but these forests have shown limited response to rising CO2 levels (Dietrich et al., 2016; Girardin et al., 2016). Carbon emissions from boreal forest soils due to increasing decomposition of organic matter and permafrost thawing might be balanced by increased productivity (Macias-Fauria et al., 2012), but boreal forests have low annual carbon gain and represent only w5% of the terrestrial carbon stocks (Fig. 2.5C). Moreover, the warming effect originating from changes in albedo due to boreal forest expansion and soil carbon emissions are expected to exert a net positive effect on global temperatures (Holden et al., 2013; Bond-Lamberty et al., 2016). It is important to note that soil carbon accumulates at very slow rates compared to land-use impacts. By one estimate, the climate change that takes place due to increases in CO2 concentration is irreversible for at least the next 1000 years, even if human-caused carbon emissions were to be completely interrupted (Solomon et al., 2009), which shows the need for maintaining and increasing carbon sinks while managing the land responsibly to prevent further emissions. Predicting the net forcing of changes in land cover is made difficult by seasonal variations in albedo, evapotranspiration, and complexities involving above-to-belowground carbon allocation, and by the industrial and agricultural emission of other greenhouse gases, such as methane and nitrous oxide (see other chapters in this book). The atmospheric levels of other pollutants, such as reactive nitrogen gases, can also lead to species- and site-specific responses to changes in climate and rising CO2 levels (Fernańdez-Martıńez et al., 2014; Silva et al., 2015d) and have blurred the boundaries between natural and managed ecosystems, causing loss of vegetation cover and soil carbon (Gruber and Galloway, 2008; Bobbink et al., 2010). One positive aspect of this realization is that managers, policy makers, and scientists now have common targets for sustainability, which include consideration of pedogenic thresholds (e.g., Vitousek and Chadwick, 2013) for increasing industrial and agricultural efficiency as well as for conserving natural ecosystems.

FINAL CONSIDERATIONS Many of the recent discoveries and promising research directions discussed in this chapter reflect collaborative contributions from natural scientists working across disciplines. The rationale for interdisciplinary integration in the natural sciences can be traced back to the notion of “land ethic,” which asserts our moral responsibility for preserving “soils, waters, plants, and animals”d collectively defined as land (Leopold, 1949). Absent from that definition is the atmosphere, which has recently become a subject of broad ethical debate in the wake of human-induced changes in climate. Interest in SPA interactions emerged from the need to include “air” in the study of terrestrial ecosystems as

48


they respond to, and exert influence on, atmospheric composition and climatic stability. Throughout this chapter, we highlighted the importance of integrative research through the lenses of ecology and biogeography, which are inherently interdisciplinary scientific fields. We also discussed scientific discoveries that will help direct future climate research that embraces biological complexity to improve predictions. Looking ahead, we will need an even broader kind of thinking, one that reconciles compartmentalized fields and that leverages diverse perspectives toward addressing the environmental challenges of the future. A strong case can be made for the integration of methods and knowledge that go beyond scientific disciplines to increase our chances of achieving environmental sustainability while ensuring food and energy security for centuries to come. Specifically, there must remain no doubt that the imminent global climate crisis is in many respects a human issue, not only because of the types of activities that drive land-use change and fossil fuel emissions but also because disciplines in the humanities and social sciences hold some of the best tools we have to understand and change human behavior. Those disciplines also provide the knowledge we need for translating scientific discoveries to broad audiences, which is a critical step toward shaping public perception and guiding the development of legislation capable of mitigating climatic impacts. Finally, it is vital to remember that the age of discovery is far from over and that public funding of research and education is essential for ensuring continued production of knowledge, which is the most promising path for moving us closer to social and environmental sustainability.

REFERENCES Ainsworth, E.A., Long, S.P., 2005. What have we learned from 15 years of free-air CO2 enrichment (FACE)? A meta-analytic review of the responses of photosynthesis, canopy. New Phytologist 165, 351e371. Alkama, R., Decharme, B., Douville, H., Ribes, A., 2011. Trends in global and basin-scale runoff over the late twentieth century: methodological issues and sources of uncertainty. Journal of Climate 24, 3000e3014. Amundson, R., Jenny, H., 1997. On a state factor model of ecosystems. BioScience 47, 536e543. Amundson, R., Sposito, G., 2013. Bridging the divide: soil resources and the geosciences on a cultivated planet. In: Geological Society of America Special Papers, vol. 501. Geological Society of America, pp. 69e80. Amundson, R., Berhe, A.A., Hopmans, J.W., et al., 2015. Soil science. Soil and human security in the 21st century. Science (New York, N.Y.) 348, 1261071. Anderson, R.G., Canadell, J.G., Randerson, J.T., et al., 2011. Biophysical considerations in forestry for climate protection. Frontiers in Ecology and the Environment 9, 174e182. Apae´stegui, J., Cruz, F.W., Sifeddine, a, et al., 2014. Hydroclimate variability of the South American Monsoon System during the last 1600 yr inferred from speleothem isotope records of the north-eastern Andes foothills in Peru. Climate of the Past Discussions 10, 533e561. Arau´jo, M.B., Rozenfeld, A., 2013. The geographic scaling of biotic interactions. Ecography 37, 406e415. Archer, D., 2010. The Global Carbon Cycle, 216 pp.

References

49

Arneth, A., Harrison, S.P., Zaehle, S., et al., 2010. Terrestrial biogeochemical feedbacks in the climate system. Nature Geoscience 3, 525e532. Asner, G.P., Knapp, D.E., Anderson, C.B., Martin, R.E., Vaughn, N., 2016a. Large-scale climatic and geophysical controls on the leaf economics spectrum. Proceedings of the National Academy of Sciences 12, E4043eE4051. Asner, G.P., Brodrick, P.G., Anderson, C.B., Vaughn, N., Knapp, D.E., Martin, R.E., 2016b. Progressive forest canopy water loss during the 2012e2015 California drought. Proceedings of the National Academy of Sciences of the United States of America 113, E249eE255. Averill, C., Turner, B.L., Finzi, A.C., 2014. Mycorrhiza-mediated competition between plants and decomposers drives soil carbon storage. Nature 505, 543e545. Bahar, N., Ishida, F., Weerasinghe, L., et al., 2016. Leaf-level photosynthetic capacity in lowland Amazonian and high elevation, Andean tropical moist forests of Peru. New Phytologist 214, 1002e1018. Baldocchi, D.D., 2005. The role of biodiversity on the evaporation of forests. In: Forest Diversity and Function. Springer-Verlag, Berlin/Heidelberg, pp. 131e148. Baldocchi, D., 2014. Biogeochemistry: managing land and climate. Nature Climate Change 4, 330e331. Barthlott, W., Hostert, A., Kier, G., et al., 2007. Geographic patterns of vascular plant diversity at continental to global scales. Erdkunde 61, 305e315. Bartlein, P.J., Harrison, S.P., Brewer, S., et al., 2011. Pollen-based continental climate reconstructions at 6 and 21 ka: a global synthesis. Climate Dynamics 37, 775e802. Beer, C., Reichstein, M., Tomelleri, E., et al., 2010. Terrestrial gross carbon dioxide uptake: global distribution and covariation with climate. Science 329, 834e838. Behling, H., Pillar, V.D., 2007. Late Quaternary vegetation, biodiversity and fire dynamics on the southern Brazilian highland and their implication for conservation and management of modern Araucaria forest and grassland ecosystems. Philosophical Transactions of the Royal Society of London. Series B, Biological Sciences 362, 243e251. Berhe, A.A., Kleber, M., 2013. Erosion, deposition, and the persistence of soil organic matter: mechanistic considerations and problems with terminology. Earth Surface Processes and Landforms 38, 908e912. Berner, R.A., 2004. The Phanerozoic Carbon Cycle : CO₂ and O₂. Oxford University Press, 150 pp. Berner, R.A., Kothavala, Z., 2001. GEOCARB III: a revised model of atmospheric CO2 over Phanerozoic time. American Journal of Science 301, 182e204. Bird, B.W., Abbott, M.B., Vuille, M., Rodbell, D.T., Stansell, N.D., Rosenmeier, M.F., 2011. A 2300-year-long annually resolved record of the South American summer monsoon from the Peruvian Andes. Proc. Natl. Acad. Sci. 108 (21), 8583e8588. Bingham, M.A., Simard, S.W., 2011. Do mycorrhizal network benefits to survival and growth of interior Douglas-fir seedlings increase with soil moisture stress? Ecology and Evolution 1, 306e316. Bingham, M.A., Simard, S., 2012. Ectomycorrhizal networks of Pseudotsuga menziesii var. glauca trees facilitate establishment of conspecific seedlings under drought. Ecosystems 15, 188e199. Blois, J.L., Zarnetske, P.L., Fitzpatrick, M.C., Finnegan, S., 2013. Climate change and the past, present, and future of biotic interactions. Science 341. Bloom, A.J., Burger, M., Rubio Asensio, J.S., Cousins, A.B., 2010. Carbon dioxide enrichment inhibits nitrate assimilation in wheat and Arabidopsis. Science (New York, N.Y.) 328, 899e903. Bobbink, R., Hicks, K., Galloway, J., et al., 2010. Global assessment of nitrogen deposition effects on terrestrial plant diversity: a synthesis. Ecological Applications 20, 30e59. Bol, R., Julich, D., Bro¨dlin, D., et al., 2016. Dissolved and colloidal phosphorus fluxes in forest ecosystems-an almost blind spot in ecosystem research. Journal of Plant Nutrition and Soil Science 79 (4), 425e438. Bonan, G.B., 2008. Forests and climate change: forcings, feedbacks, and the climate benefits of forests. Science 320, 1444e1449.

50


Bonan, G.B., Lawrence, P.J., Oleson, K.W., et al., 2011. Improving canopy processes in the Community Land Model version 4 (CLM4) using global flux fields empirically inferred from FLUXNET data. Journal of Geophysical Research 116, G02014. Bonan, G.B., Williams, M., Fisher, R.A., Oleson, K.W., 2014. Modeling stomatal conductance in the earth system: linking leaf water-use efficiency and water transport along the soileplanteatmosphere continuum. Geoscientific Model Development 7, 2193e2222. Bond, D.P.G., Wignall, P.B., 2014. Large igneous provinces and mass extinctions: an update. In: Geological Society of America Special Papers, vol. 505. Geological Society of America, pp. 29e55. Bond-Lamberty, B., Thomson, A., 2010. Temperature-associated increases in the global soil respiration record. Nature 464, 579e582. Bond-Lamberty, B., Smith, A.P., Bailey, V., 2016. Temperature and moisture effects on greenhouse gas emissions from deep active-layer boreal soils. Biogeosciences 13, 6669e6681. Boyd, A., 1992. Musopsis n. Gen.: a banana-like leaf genus from the early tertiary of eastern north Greenland. American Journal of Botany 79, 1359e1367. Bridgham, S.D., Cadillo-Quiroz, H., Keller, J.K., Zhuang, Q., 2012. Methane emissions from wetlands: biogeochemical, microbial, and modeling perspectives from local to global scales. Global Change Biology 19, 1325e1346. Brienen, R.J.W., Phillips, O.L., Feldpausch, T.R., et al., 2015. Long-term decline of the Amazon carbon sink. Nature 519, 344e348. Bristow, C.S., Hudson-Edwards, K.A., Chappell, A., 2010. Fertilizing the Amazon and equatorial Atlantic with West African dust. Geophysical Research Letters 37. Brovkin, V., Ganopolski, A., Archer, D., Munhoven, G., 2012. Glacial CO2 cycle as a succession of key physical and biogeochemical processes. Climate of the Past 8, 251e264. Brown, J.H., 2014. Why are there so many species in the tropics? Journal of Biogeography 41, 8e22. Brown, S.P., Jumpponen, A., 2015. Phylogenetic diversity analyses reveal disparity between fungal and bacterial communities during microbial primary succession. Soil Biology and Biochemistry 89, 52e60. Brundrett, M.C., 2009. Mycorrhizal associations and other means of nutrition of vascular plants: understanding the global diversity of host plants by resolving conflicting information and developing reliable means of diagnosis. Plant and Soil 320, 37e77. Carson, J.F., Whitney, B.S., Mayle, F.E., Iriarte, J., Pru¨mers, H., Soto, J.D., Watling, J., 2014. Environmental impact of geometric earthwork construction in pre-Columbian Amazonia. Proceedings of the National Academy of Sciences of the United States of America 111, 10497e10502. Ceballos, G., Ehrlich, P.R., Barnosky, A.D., Garcıá, A., Pringle, R.M., Palmer, T.M., 2015. Accelerated modern humaneinduced species losses: entering the sixth mass extinction. Science Advances 1. Chadwick, O.A., Chorover, J., 2001. The chemistry of pedogenic thresholds. Geoderma 100, 321e353. Chadwick, O.A., Derry, L.A., Vitousek, P.M., Huebert, B.J., Hedin, L.O., 1999. Changing sources of nutrients during four million years of ecosystem development. Nature 397, 491e497. Chadwick, O.A., Gavenda, R.T., Kelly, E.F., Ziegler, K., Olson, C.G., Elliott, W.C., Hendricks, D.M., 2003. The impact of climate on the biogeochemical functioning of volcanic soils. Chemical Geology 202, 195e223. Coenders-Gerrits, A.M.J., van der Ent, R.J., Bogaard, T.A., Wang-Erlandsson, L., Hrachowitz, M., Savenije, H.H.G., 2014. Uncertainties in transpiration estimates. Nature 506, E1eE2. Creed, I.F., Spargo, A.T., Jones, J.A., et al., 2014. Changing forest water yields in response to climate warming: results from long-term experimental watershed sites across North America. Global Change Biology 20, 3191e3208. Crowther, T.W., Todd-Brown, K.E.O., Rowe, C.W., et al., 2016. Quantifying global soil carbon losses in response to warming. Nature 540, 104e108.

References 51 Cruz, F.W., Vuille, M., Burns, S.J., et al., 2009. Orbitally driven eastewest antiphasing of South American precipitation. Nature Geoscience 2, 210e214. Destouni, G., Jaramillo, F., Prieto, C., 2012. Hydroclimatic shifts driven by human water use for food and energy production. Nature Climate Change 3, 213e217. Diamond, J., 2011. Collapse: How Societies Choose to Fail or Succeed, second ed. Penguin, London, UK. 589 pp. Dieleman, C.M., Branfireun, B.A., McLaughlin, J.W., Lindo, Z., 2015. Climate change drives a shift in peatland ecosystem plant community: implications for ecosystem function and stability. Global Change Biology 21, 388e395. Dietrich, R., Bell, F.W., Silva, L.C.R., Cecile, A., Horwath, W.R., Anand, M., 2016. Climatic sensitivity, water-use efficiency, and growth decline in boreal jack pine (Pinus banksiana) forests in Northern Ontario. Journal of Geophysical Research: Biogeosciences 121, 2761e2774. Dixon, J.L., Chadwick, O.A., Vitousek, P.M., 2016. Climate-driven thresholds for chemical weathering in postglacial soils of New Zealand. Journal of Geophysical Research: Earth Surface 121, 1619e1634. Doane, T.A., 2017. A survey of photogeochemistry. Geochemical Transactions 18, 1. Dokuchaev, V., 1883. Russian chernozem. In: Selected Works of VV Dokuchaev, first ed., vol. 1, pp. 14e419 (Moscow). Earles, J.M., Sperling, O., Silva, L.C.R., McElrone, A.J., Brodersen, C.R., North, M.P., Zwieniecki, M.A., 2016. Bark water uptake promotes localized hydraulic recovery in coastal redwood crown. Plant, Cell and Environment 39, 320e328. Edwards, E.J., Still, C.J., 2008. Climate, phylogeny and the ecological distribution of C4 grasses. Ecology Letters 11, 266e276. Eller, C.B., Lima, A.L., Oliveira, R.S., 2013. Foliar uptake of fog water and transport belowground alleviates drought effects in the cloud forest tree species, Drimys brasiliensis (Winteraceae). New Phytologist 199, 151e162. Elmore, A.J., Nelson, D.M., Craine, J.M., et al., 2016. Earlier springs are causing reduced nitrogen availability in North American eastern deciduous forests. Nature Plants 2, 16133. Enquist, B.J., Norberg, J., Bonser, S.P., et al., 2015. Scaling from traits to ecosystems: developing a general Trait Driver Theory via integrating trait-based and metabolic scaling theories. Advances in Ecological Research 52, 249e318. van der Ent, A., Lambers, H., 2016. Plant-soil interactions in global biodiversity hotspots. Plant and Soil 403, 1e5. Epica, C.M., 2004. Eight glacial cycles from an Antarctic ice core. Nature 429, 623e628. Evans, M.E.K., Merow, C., Record, S., McMahon, S.M., Enquist, B.J., 2016. Towards process-based range modeling of many species. Trends in Ecology & Evolution 31, 860e871. Evaristo, J., Jasechko, S., McDonnell, J.J., 2015. Global separation of plant transpiration from groundwater and streamflow. Nature 525, 91e94. Fernandez, C.W., Nguyen, N.H., Stefanski, A., et al., 2016. Ectomycorrhizal fungal response to warming is linked to poor host performance at the boreal-temperate ecotone. Global Change Biology 23, 1598e1609. Fernańdez-Martıńez, M., Vicca, S., Janssens, I.A., et al., 2014. Nutrient availability as the key regulator of global forest carbon balance. Nature Climate Change 4, 471e476. Field, C.B., Lobell, D.B., Peters, H.A., Chiariello, N.R., 2007. Feedbacks of terrestrial ecosystems to climate change. Annual Review of Environment and Resources 32, 1e29. Field, K.J., Pressel, S., Duckett, J.G., Rimington, W.R., Bidartondo, M.I., 2015. Symbiotic options for the conquest of land. Trends in Ecology & Evolution 30, 477e486. Fine, P.V.A., 2015. Ecological and evolutionary drivers of geographic variation in species diversity. Annual Review of Ecology Evolution and Systematics 46, 369e392. Finzi, A.C., Moore, D.J.P., DeLucia, E.H., et al., 2006. Progressive nitrogen limitation of ecosystem processes under elevated CO2 in a warm-temperate forest. Ecology 87, 15e25. Fluckiger, J., Monnin, E., Stauffer, B., et al., 2002. High-resolution Holocene N2O ice core record and its relationship with CH4 and CO2. Global Biogeochemical Cycles 16 art. no.-1010 ST-High-resolution Holocene N2O i.

52


Food and Agriculture Organization, 2012. World food and agriculture in review. In: The State of Food and Agriculture, pp. 97e106. Franco, A.C., Rossatto, D.R., de Carvalho Ramos Silva, L., da Silva Ferreira, C., 2014. Cerrado vegetation and global change: the role of functional types, resource availability and disturbance in regulating plant community responses to rising CO2 levels and climate warming. Theoretical and Experimental Plant Physiology 26, 19e38. Friedlingstein, P., Meinshausen, M., Arora, V.K., et al., 2014. Uncertainties in CMIP5 climate projections due to carbon cycle feedbacks. Journal of Climate 27, 511e526. Friis, E.M., Crane, P.R., Pedersen, K.R., 2011. Early Flowers and Angiosperm Evolution. Cambridge University Press, Cambridge, UK, 585 pp. Friis, E.M., Pedersen, K.R., Crane, P.R., 2016. The emergence of core eudicots: new floral evidence from the earliest Late Cretaceous. Proceedings of the Royal Society of London. Series B, Biological Sciences 283. Fritz, S.C., Anderson, N.J., 2013. The relative influences of climate and catchment processes on Holocene lake development in glaciated regions. Journal of Paleolimnology 49, 349e362. Go´mez-Guerrero, A., Silva, L.C.R., Barrera-Reyes, M., Kishchuk, B., VelazquezMartinez, A., Martinez-Trinidad, T., Ofelia Plascencia-Escalante, F., Horwath, W.R., 2013. Growth decline and divergent tree ring isotopic composition (13C and 18O) contradict predictions of CO2 stimulation in high altitudinal forests. Global Change Biol. 19 (6), 1748e1758. Girardin, M.P., Bouriaud, O., Hogg, E.H., et al., 2016. No growth stimulation of Canada’s boreal forest under half-century of combined warming and CO2 fertilization. Proceedings of the National Academy of Sciences of the United States of America 113, E8406eE8414. Gorzelak, M.A., Asay, A.K., Pickles, B.J., Simard, S.W., 2015. Inter-plant communication through mycorrhizal networks mediates complex adaptive behaviour in plant communities. AoB Plants 7, 2041e2851. Grau, O., Pen˜uelas, J., Ferry, B., et al., 2017. Nutrient-cycling mechanisms other than the direct absorption from soil may control forest structure and dynamics in poor Amazonian soils. Scientific Reports 7, 45017. Gross, A., Turner, B.L., Goren, T., Berry, A., Angert, A., 2016. Tracing the sources of atmospheric phosphorus deposition to a tropical rain forest in Panama using stable oxygen isotopes. Environmental Science and Technology 50, 1147e1156. Gruber, N., Galloway, J.N., 2008. An Earth-system perspective of the global nitrogen cycle. Nature 451, 293e296. Guilherme Pereira, C., Clode, P.L., Oliveira, R.S., Lambers, H., 2018. Eudicots from severely phosphorus-impoverished environments preferentially allocate phosphorus to their mesophyll. New Phytol 218, 959e973. Hahm, W.J., Riebe, C.S., Lukens, C.E., Araki, S., 2014. Bedrock composition regulates mountain ecosystems and landscape evolution. Proceedings of the National Academy of Sciences of the United States of America 111, 3338e3343. Haridasan, M., 2008. Nutritional adaptations of native plants of the cerrado biome in acid soils. Brazilian Journal of Plant Physiology 20, 183e195. Harrison, S., Damschen, E., Fernandez-Going, B., Eskelinen, A., Copeland, S., 2015. Plant communities on infertile soils are less sensitive to climate change. Annals of Botany 116, 1017e1022. Harrison, S.P., Bartlein, P.J., Prentice, I.C., 2016. What have we learnt from palaeoclimate simulations? Journal of Quaternary Science 31, 363e385. Hayes, P., Turner, B.L., Lambers, H., Laliberte´, E., 2014. Foliar nutrient concentrations and resorption efficiency in plants of contrasting nutrient-acquisition strategies along a 2million-year dune chronosequence. J Ecol 102, 396e410. He, Y., Trumbore, S.E., Torn, M.S., Harden, J.W., Vaughn, L.J.S., Allison, S.D., Randerson, J.T., 2016. Radiocarbon constraints imply reduced carbon uptake by soils during the 21st century. Science 353, 1419e1424.

References

53

Higgins, S.I., 2017. Ecosystem assembly: a mission for terrestrial earth system science. Ecosystems 20, 69e77. Hinsinger, P., 2001. Bioavailability of soil inorganic P in the rhizosphere as affected by rootinduced chemical changes: a review. Plant and Soil 237, 173e195. Hoffmann, W.A., Geiger, E.L., Gotsch, S.G., et al., 2012. Ecological thresholds at the savanna-forest boundary: how plant traits, resources and fire govern the distribution of tropical biomes. Ecology Letters 15, 759e768. Holden, S.R., Gutierrez, A., Treseder, K.K., 2013. Changes in soil fungal communities, extracellular enzyme activities, and litter decomposition across a fire chronosequence in Alaskan boreal forests. Ecosystems 16, 34e46. Hopper, S.D., 2009. OCBIL theory: towards an integrated understanding of the evolution, ecology and conservation of biodiversity on old, climatically buffered, infertile landscapes. Plant and Soil 322, 49e86. Houghton, R.A., 2007. Balancing the global carbon budget. Annual Review of Earth and Planetary Sciences 35, 313e347. Houlton, B.Z., Wang, Y.-P., Vitousek, P.M., Field, C.B., 2008. A unifying framework for dinitrogen fixation in the terrestrial biosphere. Nature 454, 327e330. IPCC, 2013. Summary for policymakers. In: Stocker, T.F., Qin, D., Plattner, G.-K., Tignor, M., Allen, S.K., Boschung, J., Nauels, A., Xia, Y., Bex, V., Midgley, P.M. (Eds.), Climate Change 2013: The Physical Science Basis. Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA, p. 27. Izquierdo, J.E., Houlton, B.Z., van Huysen, T.L., 2013. Evidence for progressive phosphorus limitation over long-term ecosystem development: examination of a biogeochemical paradigm. Plant and Soil 367, 135e147. Jasechko, S., Sharp, Z.D., Gibson, J.J., Birks, S.J., Yi, Y., Fawcett, P.J., 2013. Terrestrial water fluxes dominated by transpiration. Nature 496, 347e350. Jenny, H., 1941. Factors of soil formation. Soil Science 52, 415. Jerszurki, D., Souza, J.L.M., Silva, L.C.R., 2017. Expanding the geography of evapotranspiration: an improved method to quantify land-to-air water fluxes in tropical and subtropical regions. PLoS One 12, e0180055. Joosten, H., 2015. Current soil carbon loss and land degradation globally: where are the hotspots and why there? Soil Carbon: Science, Management and Policy for Multiple Benefits 71, 224e234. Kallenbach, C.M., Frey, S.D., Grandy, A.S., et al., 2016. Direct evidence for microbialderived soil organic matter formation and its ecophysiological controls. Nature Communications 7, 13630. Keenan, T.F., Hollinger, D.Y., Bohrer, G., Dragoni, D., Munger, J.W., Schmid, H.P., Richardson, A.D., 2013. Increase in forest water-use efficiency as atmospheric carbon dioxide concentrations rise. Nature 499, 324e327. Kenrick, P., Strullu-Derrien, C., 2014. The origin and early evolution of roots. Plant Physiology 166, 570e580. Kier, G., Kreft, H., Lee, T.M., et al., 2009. A global assessment of endemism and species richness across island and mainland regions. Proceedings of the National Academy of Sciences of the United States of America 106, 9322e9327. Ko¨rner, C., 2012. Alpine Treelines. Springer, Basel, 220 pp. Ko¨rner, C., 2017. A matter of tree longevity. Science 355, 130e131. Ladd, B., Laffan, S.W., Amelung, W., et al., 2012. Estimates of soil carbon concentration in tropical and temperate forest and woodland from available GIS data on three continents. In: Wright, I. (Ed.), Global Ecology and Biogeography, vol. 22, pp. 461e469. Lal, R., 2008. Carbon sequestration. Philosophical Transactions of the Royal Society of London. Series B, Biological Sciences 363, 815e830. Laliberte´, E., Grace, J.B., Huston, M.A., Lambers, H., Teste, F.P., Turner, B.L., Wardle, D.A., 2013. How does pedogenesis drive plant diversity? Trends in Ecology & Evolution 28, 331e340.

54


Lambers, H., Raven, J.A., Shaver, G.R., Smith, S.E., 2008. Plant nutrient-acquisition strategies change with soil age. Trends in Ecology & Evolution 23, 95e103. Lambers, H., Mougel, C., Jaillard, B., Hinsinger, P., 2009. Plant-microbe-soil interactions in the rhizosphere: an evolutionary perspective. Plant and Soil 321, 83e115. Lambers, H., Brundrett, M.C., Raven, J.A., Hopper, S.D., 2010. Plant mineral nutrition in ancient landscapes: high plant species diversity on infertile soils is linked to functional diversity for nutritional strategies. Plant and Soil 348, 11e31. Lambers, H., Martinoia, E., Renton, M., 2015. Plant adaptations to severely phosphorusimpoverished soils. Current Opinion in Plant Biology 25, 23e31. Lambers, H., Albornoz, F., Kotula, L., Laliberte´, E., Ranathunge, K., Teste, F.P., Zemunik, G., 2018. How belowground interactions contribute to the coexistence of mycorrhizal and non-mycorrhizal species in severely phosphorus-impoverished hyperdiverse ecosystems. Plant Soil 424, 11e34. Lamy, F., Kaiser, J., 2009. Glacial to Holocene paleoceanographic and continental paleoclimate reconstructions based on ODP Site 1233/GeoB 3313 off southern Chile. In: Vimeux, F., Sylvestre, F., Khodri, M. (Eds.), Past Climate Variability in South America and Surrounding Regions Developments in Paleoenvironmental Research 14, vol. 14. Springer, pp. 129e156. Law, B.E., Falge, E., Gu, L., Baldocchi, D.D., Bakwin, P., 2002. Environmental controls over carbon dioxide and water vapor exchange of terrestrial vegetation. Agricultural and Forest Meteorology 113, 97e120. Ledru, M.-P., 2002. Late quaternary history and evolution of the cerrados as revealed by palynological records. In: Oliveira, P.S., Marquis, R.J. (Eds.), The Cerrados of Brazil Ecology and Natural History of a Neotropical Savanna. Columbia University Press, pp. 33e50. Leopold, A., 1949. The Land Ethic. A Sand Country Almanac. Limm, E.B., Simonin, K.A., Bothman, A.G., Dawson, T.E., 2009. Foliar water uptake: a common water acquisition strategy for plants of the redwood forest. Oecologia 161, 449e459. Loftus, P.J., Cohen, A.M., Long, J.C.S., Jenkins, J.D., 2015. A critical review of global decarbonization scenarios: what do they tell us about feasibility? Wiley Interdisciplinary Reviews: Climate Change 6, 93e112. Lu¨thi, D., Le Floch, M., Bereiter, B., et al., 2008. High-resolution carbon dioxide concentration record 650,000-800,000 years before present. Nature 453, 379e382. Macias-Fauria, M., Forbes, B.C., Zetterberg, P., Kumpula, T., 2012. Eurasian Arctic greening reveals teleconnections and the potential for structurally novel ecosystems. Nature Climate Change 2, 1e6. Mankin, J.S., Smerdon, J.E., Cook, B.I., et al., 2017. The curious case of projected 21st-century drying but greening in the American West. Journal of Climate. JCLI-D-17-0213.1. 30 (21), 8689e8710. Marchant, R., Almeida, L., Behling, H., et al., 2002. Distribution and ecology of parent taxa of pollen lodged within the Latin American Pollen Database. Review of Palaeobotany and Palynology 121, 1e75. Maxwell, T., Silva, L.C.R., Horwath, W.R., 2014. Using multielement isotopic analysis to decipher drought impacts and adaptive management in ancient agricultural systems. Proceedings of the National Academy of Sciences of the United States of America 111, E4807eE4808. Maxwell, T.M., Silva, L.C.R., Horwath, W.R., 2018a. Integrating effects of species composition and soil properties to predict shifts in montane forest carbon-water relations PNAS, 115 (18), E4219eE4226. Maxwell, T.M., Silva, L.C.R., Horwath, W.R., 2018b. (In Press): Predictable oxygen isotope exchange between plant lipids and Environmental water: implications for ecosystem water balance reconstruction First published: 2018. https://doi.org/10.1029/ 2018JG004553. Medlyn, B.E., Zaehle, S., De Kauwe, M.G., et al., 2015. Using ecosystem experiments to improve vegetation models. Perspective 5, 528e534.

References 55 Meyer, K.M., Klein, A.M., Rodrigues, J.L.M., et al., 2017. Conversion of Amazon rainforest to agriculture alters community traits of methane-cycling organisms. Molecular Ecology 26, 1547e1556. Mittermeier, R.A., Turner, W.R., Larsen, F.W., Brooks, T.M., Gascon, C., 2011. Global biodiversity conservation: the critical role of hotspots. In: Biodiversity Hotspots Distribution and Protection of Conservation Priority Areas. Springer, New York, NY, pp. 3e22. Monson, R., Baldocchi, D., 2014. Terrestrial Biosphere-Atmosphere Fluxes. Cambridge University Press, Cambridge, 507 pp. Montan˜ez, I.P., Mcelwain, J.C., Poulsen, C.J., et al., 2016. Climate, pCO2 and terrestrial carbon cycle linkages during late Palaeozoic glacialeinterglacial cycles. Nature Geoscience 9, 824e828. Morford, S.L., Houlton, B.Z., Dahlgren, R.A., 2011. Increased forest ecosystem carbon and nitrogen storage from nitrogen rich bedrock. Nature 477, 78e81. Morford, S.L., Houlton, B.Z., Dahlgren, R.A., 2016. Geochemical and tectonic uplift controls on rock nitrogen inputs across terrestrial ecosystems. Global Biogeochemical Cycles 30, 333e349. Morris, J., Ye, R., Silva, L.C.R., Horwath, W.R., 2017. Nitrogen fertilization had no effect on CH and NO emissions in rice planted in rewetted peatlands. Soil Science Society of America Journal 81, 224. Moulton, K.L., West, J., Berner, R.A., 2000. Solute flux and mineral mass balance approaches to the quantification of plant effects on silicate weathering. American Journal of Science 300, 539e570. Myers, N., Mittermeier, R a, Mittermeier, C.G., da Fonseca, G a, Kent, J., 2000. Biodiversity hotspots for conservation priorities. Nature 403, 853e858. Myers, S.S., Zanobetti, A., Kloog, I., et al., 2014. Increasing CO2 threatens human nutrition. Nature 510, 139e142. Newbold, T., Hudson, L.N., Arnell, A.P., et al., 2016. Has land use pushed terrestrial biodiversity beyond the planetary boundary? A global assessment. Science 353, 288e291. Nobre, C.A., Sampaio, G., Borma, L.S., Castilla-Rubio, J.C., Silva, J.S., Cardoso, M., 2016. Land-use and climate change risks in the Amazon and the need of a novel sustainable development paradigm. Proceedings of the National Academy of Sciences of the United States of America 113, 10759e10768. Nova´k, V., 2012. Evapotranspiration in the Soil-plant-Atmosphere System. Springer, Netherlands, Dordrecht. Oelkers, E.H., Cole, D.R., 2008. Carbon dioxide sequestration a solution to a global problem. Elements 4. Oliveira, R.S., Dawson, T.E., Burgess, S.S.O., 2005. Evidence for direct water absorption by the shoot of the desiccation-tolerant plant Vellozia flavicans in the savannas of central Brazil. Journal of Tropical Ecology 21, 585e588. Oliveira, R.S., Galvaõ, H.C., de Campos, M.C.R., Eller, C.B., Pearse, S.J., Lambers, H., 2015. Mineral nutrition of campos rupestres plant species on contrasting nutrient-impoverished soil types. New Phytologist 205, 1183e1194. Paiva, A.O., Silva, L.C.R., Haridasan, M., 2015. Productivity-efficiency tradeoffs in tropical gallery forest-savanna transitions: linking plant and soil processes through litter input and composition. Plant Ecology 216, 775e787. Parikh, S.J., Mukome, F.N.D., Zhang, X., 2014. ATReFTIR spectroscopic evidence for biomolecular phosphorus and carboxyl groups facilitating bacterial adhesion to iron oxides. Colloids and Surfaces B: Biointerfaces 119, 38e46. Pecl, G.T., Arau´jo, M.B., Bell, J.D., et al., 2017. Biodiversity redistribution under climate change: impacts on ecosystems and human well-being. Science 355. Pe´rez, C.A., Carmona, M.R., Aravena, J.C., Armesto, J.J., 2004. Successional changes in soil nitrogen availability, non-symbiotic nitrogen fixation and carbon/nitrogen ratios in southern Chilean forest ecosystems. Oecologia 140, 617e625. Pessenda, L.C.R., De Oliveira, P.E., Mofatto, M., et al., 2009. The evolution of a tropical rainforest/grassland mosaic in southeastern Brazil since 28,000 14C yr BP based on carbon isotopes and pollen records. Quaternary Research 71, 437e452.

56


Pfeifer-Meister, L., Bridgham, S.D., Reynolds, L.L., et al., 2016. Climate change alters plant biogeography in Mediterranean prairies along the West Coast, USA. Global Change Biology 22, 845e855. Philip, J.R., 1966. Plant water relations: some physical aspects. Annual Review of Plant Physiology 17, 245e268. Pinter, N., Fiedel, S., Keeley, J.E., 2011. Fire and vegetation shifts in the Americas at the vanguard of Paleoindian migration. Quaternary Science Reviews 30, 269e272. Polgar, C.A., Primack, R.B., 2011. Leaf-out phenology of temperate woody plants: from trees to ecosystems. New Phytologist 191, 926e941. Potter, C., Boriah, S., Steinbach, M., Kumar, V., Klooster, S., 2008. Terrestrial vegetation dynamics and global climate controls. Climate Dynamics 31, 67e78. Prentice, I.C., Harrison, S.P., Bartlein, P.J., 2011. Global vegetation and terrestrial carbon cycle changes after the last ice age. New Phytologist 189, 988e998. Qiao, Y., Miao, S., Silva, L.C.R., Horwath, W.R., 2014. Understory species regulate litter decomposition and accumulation of C and N in forest soils: a long-term dual-isotope experiment. Forest Ecology and Management 329, 318e327. Que´re´, C.Le, et al., 2016. Global carbon budget 2016. Earth Systems Science Data 8, 605e649. Rasmussen, C., Southard, R.J., Horwath, W.R., 2008. Litter type and soil minerals control temperate forest soil carbon response to climate change. Global Change Biology 14, 2064e2080. Raven, J.A., Edwards, D., 2001. Roots: evolutionary origins and biogeochemical significance. Journal of Experimental Botany 52, 381e401. Raven, J.A., Lambers, H., Smith, S.E., Westoby, M., 2018. Costs of acquiring phosphorus by vascular land plants: patterns and implications for plant coexistence. New Phytol 217, 1420e1427. Read, D.J., Perez-Moreno, J., 2003. Mycorrhizas and nutrient cycling in ecosystems - a journey towards relevance? New Phytologist 157, 475e492. Reed, S.C., Cleveland, C.C., Townsend, A.R., 2011. Functional ecology of free-living nitrogen fixation: a contemporary perspective. Annual Review of Ecology Evolution and Systematics 42, 489e512. Reich, P.B., 2012. Key canopy traits drive forest productivity. Proceedings. Biological Sciences/The Royal Society 279, 2128e2134. Reich, P.B., 2014. The world-wide “fasteslow” plant economics spectrum: a traits manifesto. Journal of Ecology 102, 275e301. Reich, P.B., Hobbie, S.E., Lee, T.D., 2014. Plant growth enhancement by elevated CO2 eliminated by joint water and nitrogen limitation. Nature Geoscience 7, 920e924. Riehl, S., Pustovoytov, K.E., Weippert, H., Klett, S., Hole, F., 2014. Drought stress variability in ancient Near Eastern agricultural systems evidenced by d13C in barley grain. Proceedings of the National Academy of Sciences of the United States of America 111, 12348e12353. Robin, A., Vansuyt, G., Hinsinger, P., Meyer, J.M., Briat, J.F., Lemanceau, P., 2008. Chapter 4 iron dynamics in the rhizosphere: consequences for plant health and nutrition. Advances in Agronomy 99, 183e225. Rodrigues, J.L.M., Pellizari, V.H., Mueller, R., et al., 2013. Conversion of the Amazon rainforest to agriculture results in biotic homogenization of soil bacterial communities. Proceedings of the National Academy of Sciences of the United States of America 110, 988e993. Rohde, R.A., Muller, R.A., 2005. Cycles in fossil diversity. Nature 434, 208e210. Rossatto, D.R., Hoffmann, W.A., Franco, A.C., 2009. Differences in growth patterns between co-occurring forest and savanna trees affect the forest-savanna boundary. Functional Ecology 23, 689e698. Royer, D.L., 2006. CO2-forced climate thresholds during the Phanerozoic. Geochimica et Cosmochimica Acta 70, 5665e5675. Royer, D.L., Berner, R.A., Montanez, I.P., Tabor, N.J., Beerling, D.J., 2004. CO2 as a primary driver of Phanerozoic climate. Geological Society of America Today.

References 57 Sachse, D., Billault, I., Bowen, G.J., et al., 2012. Molecular paleohydrology: interpreting the hydrogen-isotopic composition of lipid biomarkers from photosynthesizing organisms. Annual Review of Earth and Planetary Sciences 40, 221e249. Saha, A.K., Setegn, S.G., 2015. Ecohydrology: understanding and maintaining ecosystem services for IWRM. In: Sustainability of Integrated Water Resources Management. Springer International Publishing, Cham, pp. 121e145. Salzer, M.W., Hughes, M.K., Bunn, A.G., Kipfmueller, K.F., 2009. Recent unprecedented tree-ring growth in bristlecone pine at the highest elevations and possible causes. Proceedings of the National Academy of Sciences of the United States of America 106, 20348e20353. Schimel, D.S., Asner, G.P., Moorcroft, P., 2013. Observing changing ecological diversity in the Anthropocene. Frontiers in Ecology and the Environment. 130118053603002. 11 (3), 129e137. Schlesinger, W.H., Jasechko, S., 2014. Transpiration in the global water cycle. Agricultural and Forest Meteorology 189e190, 115e117. Schulte, P., Alegret, L., Arenillas, I., et al., 2010. The chicxulub asteroid impact and mass extinction at the cretaceous-paleogene boundary. Science 327, 1214e1218. Shaw, S.B., Mehta, D., Riha, S.J., 2014. Using simple data experiments to explore the influence of non-temperature controls on maize yields in the mid-West and Great Plains. Climatic Change 122, 747e755. Sheil, D., Ladd, B., Silva, L.C.R., Laffan, S.W., Van Heist, M., 2016. How are soil carbon and tropical biodiversity related? Environmental Conservation 1e11. Shi, C., Silva, L.C.R., Zhang, H., Zheng, Q., Xiao, B., Wu, N., Sun, G., 2015. Climate warming alters nitrogen dynamics and total non-structural carbohydrate accumulations of perennial herbs of distinctive functional groups during the plant senescence in autumn in an alpine meadow of the Tibetan Plateau, China. Agricultural and Forest Meteorology 200, 21e29. Silva, L.C.R., Anand, M., 2011. Mechanisms of Araucaria (Atlantic) forest expansion into southern Brazilian grasslands. Ecosystems 14, 1354e1371. Silva, L.C.R., 2014. Importance of climate-driven forestesavanna biome shifts in anthropological and ecological research. Proceedings of the National Academy of Sciences of the United States of America 111, E3831eE3832. Silva, L.C.R., 2015. From air to land: understanding water resources through plant-based multidisciplinary research. Trends in Plant Science 20, 399e401. Silva, L.C.R., 2017. Carbon sequestration beyond tree longevity. Science 355, 1141. Silva, L.C.R., Anand, M., 2013. Historical links and new frontiers in the study of forestatmosphere interactions. Community Ecology 14, 208e218. Silva, L.C.R., Sternberg, L., Haridasan, M., Hoffmann, W.A., Miralles-Wilhelm, F., Franco, A.C., 2008. Expansion of gallery forests into central Brazilian savannas. Global Change Biology 14, 2108e2118. Silva, L.C.R., Haridasan, M., Hoffmann, W.A., 2010. Not all forests are expanding over central Brazilian savannas. Plant and Soil 333, 431e442. Silva, L.C.R., Salamanca-Jimenez, A., Doane, T.A., Horwath, W.R., 2015a. Carbon dioxide level and form of soil nitrogen regulate assimilation of atmospheric ammonia in young trees. Scientific Reports 5, 13141. Silva, L.C.R., Doane, T.A., Correâ, R.S., Valverde, V., Pereira, E.I.P., Horwath, W.R., 2015b. Iron-mediated stabilization of soil carbon amplifies the benefits of ecological restoration in degraded lands. Ecological Applications 25, 1226e1234. Silva, L.C.R., Pedroso, G., Doane, T.A., Mukome, F.N.D., Horwath, W., 2015c. Beyond the cellulose: oxygen isotope composition of plant lipids as a proxy for terrestrial water balance. Geochemical Perspective Letters 1, 33e42. Silva, L.C.R., Go´mez-Guerrero, A., Doane, T.A., Horwath, W.R., 2015d. Isotopic and nutritional evidence for species- and site-specific responses to N deposition and rising atmospheric CO2 in temperate forests. Journal of Geophysical Research: Biogeosciences 120, 1110e1123.

58


Silva, L.C.R., Sun, G., Zhu-Barker, X., Liang, Q., Wu, N., Horwath, W.R., 2016. Tree growth acceleration and expansion of alpine forests: the synergistic effect of atmospheric and edaphic change. Science Advances 2, e1501302. Silveira, F.A.O., Negreiros, D., Barbosa, N.P.U., et al., 2016. Ecology and evolution of plant diversity in the endangered campo rupestre: a neglected conservation priority. Plant and Soil 403, 129e152. Simard, S.W., 2009. The foundational role of mycorrhizal networks in self-organization of interior Douglas-fir forests. Forest Ecology and Management 258, S95eS107. Simard, S.W., Beiler, K.J., Bingham, M.A., Deslippe, J.R., Philip, L.J., Teste, F.P., 2012. Mycorrhizal networks: mechanisms, ecology and modelling. Fungal Biology Reviews 26, 39e60. van der Sleen, P., Groenendijk, P., Vlam, M., et al., 2014. No growth stimulation of tropical trees by 150 years of CO2 fertilization but water-use efficiency increased. Nature Geoscience 8, 24e28. Smith, S.E., Read, D.J., David, J., 2008. Mycorrhizal Symbiosis. Academic Press, 787 pp. Smith, S.E., Anderson, I.C., Smith, F.A., 2015. Mycorrhizal associations and phosphorus acquisition: from cells to ecosystems. In: Annual Plant Reviews, vol. 48. John Wiley & Sons, Inc., Hoboken, N.J., USA, pp. 409e439. Smith, W.K., Reed, S.C., Cleveland, C.C., et al., 2016. Large divergence of satellite and Earth system model estimates of global terrestrial CO2 fertilization. Nature Climate Change 6, 306e310. Solomon, S., Plattner, G.-K., Knutti, R., Friedlingstein, P., 2009. Irreversible climate change due to carbon dioxide emissions. Proceedings of the National Academy of Sciences of the United States of America 106, 1704e1709. Song, Y.Y., Simard, S.W., Carroll, A., Mohn, W.W., Zeng, R.S., 2015. Defoliation of interior Douglas-fir elicits carbon transfer and stress signalling to ponderosa pine neighbors through ectomycorrhizal networks. Scientific Reports 5, 8495. Sperling, O., Silva, L.C.R., Tixier, A., The´roux-Rancourt, G., Zwieniecki, M.A., 2017. Temperature gradients assist carbohydrate allocation within trees. Scientific Reports 7, 3265. Staal, A., Dekker, S.C., Xu, C., van Nes, E.H., 2016. Bistability, spatial interaction, and the distribution of tropical forests and savannas. Ecosystems 19, 1080e1091. Stevens, N., Lehmann, C.E.R., Murphy, B.P., Durigan, G., 2016. Savanna woody encroachment is widespread across three continents. Global Change Biology 23 (1), 235e244. Sullivan, B.W., Smith, W.K., Townsend, A.R., Nasto, M.K., Reed, S.C., Chazdon, R.L., Cleveland, C.C., 2014. Spatially robust estimates of biological nitrogen (N) fixation imply substantial human alteration of the tropical N cycle. Proceedings of the National Academy of Sciences of the United States of America 111, 8101e8106. Sullivan, M.J.P., Talbot, J., Lewis, S.L., et al., 2017. Diversity and carbon storage across the tropical forest biome. Scientific Reports 7, 39102. Svenning, J.-C., Eiserhardt, W.L., Normand, S., Ordonez, A., Sandel, B., 2015. The influence of paleoclimate on present-day patterns in biodiversity and ecosystems. Annual Review of Ecology Evolution and Systematics 46, 551e572. Taylor, L.L., Leake, J.R., Quirk, J., Hardy, K., Banwart, S.A., Beerling, D.J., 2009. Biological weathering and the long-term carbon cycle: integrating mycorrhizal evolution and function into the current paradigm. Geobiology 7, 171e191. Taylor, Z.P., Horn, S.P., Mora, C.I., Orvis, K.H., Cooper, L.W., 2010. A multi-proxy palaeoecological record of late-Holocene forest expansion in lowland Bolivia. Palaeogeography, Palaeoclimatology, Palaeoecology 293, 98e107. Taylor, L.L., Banwart, S.A., Valdes, P.J., Leake, J.R., Beerling, D.J., 2012. Evaluating the effects of terrestrial ecosystems, climate and carbon dioxide on weathering over geological time: a global-scale process-based approach. Philosophical Transactions of the Royal Society of London. Series B, Biological Sciences 367, 565e582. Terrer, C., Vicca, S., Hungate, B.A., et al., 2016. Mycorrhizal association as a primary control of the CO₂ fertilization effect. Science (New York, N.Y.) 353, 72e74.

References 59 Teste, F.P., Simard, S.W., Durall, D.M., Guy, R.D., Berch, S.M., 2010. Net carbon transfer between Pseudotsuga menziesii var. glauca seedlings in the field is influenced by soil disturbance. Journal of Ecology 98, 429e439. Thomas, C.D., Cameron, A., Green, R.E., et al., 2004. Extinction risk from climate change. Nature 427, 145e148. Trivedi, P., Delgado-Baquerizo, M., Trivedi, C., et al., 2016. Microbial regulation of the soil carbon cycle: evidence from geneeenzyme relationships. The ISME Journal 10, 2593e2604. Trumbore, S., 2009. Radiocarbon and soil carbon dynamics. Annual Review of Earth and Planetary Sciences 37, 47e66. Tuthorn, M., Zech, M., Ruppenthal, M., et al., 2014. Oxygen isotope ratios (18O/16O) of hemicellulose-derived sugar biomarkers in plants, soils and sediments as paleoclimate proxy II: insight from a climate transect study. Geochimica et Cosmochimica Acta 126, 624e634. Vandermark, D., Tarduno, J.A., Brinkman, D.B., Cottrell, R.D., Mason, S., 2009. New late cretaceous macrobaenid turtle with Asian affinities from the high canadian Arctic: dispersal via ice-free polar routes. Geology 37, 183e186. Veizer, J., Godderis, Y., François, L.M., 2000. Evidence for decoupling of atmospheric CO2 and global climate during the Phanerozoic eon. Nature 408, 698e701. Vitousek, P.M., 2004. Nutrient Cycling and Limitation: Hawai’i as a Model System, vol. 30. Princeton University Press, 223 pp. Vitousek, P.M., Chadwick, O.A., 2013. Pedogenic thresholds and soil process domains in basalt-derived soils. Ecosystems 16, 1379e1395. Vitousek, P.M., Porder, S., Houlton, B.Z., Chadwick, O.A., 2010. Terrestrial phosphorus limitation: mechanisms, implications, and nitrogenephosphorus interactions. Ecological Applications 20, 5e15. Walker, T.W., Syers, J.K., 1976. The fate of phosphorus during pedogenesis. Geoderma 15, 1e19. Wang, Z., Hoomann, T., Six, J., Kaplan, J.O., Govers, G., Doetterl, S., Oost, K.Van, 2017. Human-induced erosion has offset one-third of carbon emissions from land cover change. Nature Climate Change 7, 345e349. Weremijewicz, Joanna; Sternberg, Leonel da Silveira Lobo O’Reilly; Janos, David P., 2016. Common mycorrhizal networks amplify competition by preferential mineral nutrient allocation to large host plants. New Phytologist 212(2), 461e471. Whiteside, J.H., Grice, K., 2016. Biomarker records associated with mass extinction events. Annual Review of Earth and Planetary Sciences 44, 581e612. Whittaker, R.H., 1975. Communities and Ecosystems. Collier MacMillan, London, UK, 352 pp. Wieder, W.R., Allison, S.D., Davidson, E.A., et al., 2015. Explicitly representing soil microbial processes in Earth system models. Global Biogeochemical Cycles 29, 1782e1800. Williams, K.J., Ford, A., Rosauer, D.F., et al., 2011. Forests of east Australia: the 35th biodiversity hotspot. In: Biodiversity Hotspots, pp. 295e310. Williams, A.P., Allen, C.D., Macalady, A.K., et al., 2012a. Temperature as a potent driver of regional forest drought stress and tree mortality. Nature Climate Change 3, 292e297. Williams, C.A., Reichstein, M., Buchmann, N., et al., 2012b. Climate and vegetation controls on the surface water balance: synthesis of evapotranspiration measured across a global network of flux towers. Water Resources Research 48, W06523. Winsome, T., Silva, L.C.R., Scow, K.M., Doane, T.A., Powers, R.F., Horwath, W.R., 2017. Plant-microbe interactions regulate carbon and nitrogen accumulation in forest soils. Forest Ecology and Management 384, 415e423. Wortham, B.E., Wong, C.I., Silva, L.C.R., et al., 2017. Assessing response of local moisture conditions in central Brazil to variability in regional monsoon intensity using speleothem 87Sr/86Sr values. Earth and Planetary Science Letters 463, 310e322. Wright, I.J., Reich, P.B., Westoby, M., et al., 2004. The worldwide leaf economics spectrum. Nature 428, 821e827.

60


Yospin, G.I., Bridgham, S.D., Neilson, R.P., et al., 2015. A new model to simulate climatechange impacts on forest succession for local land management. Ecological Applications 25, 226e242. Yu, H., Chin, M., Yuan, T., et al., 2015. The fertilizing role of African dust in the Amazon rainforest: a first multiyear assessment based on data from Cloud-Aerosol Lidar and Infrared Pathfinder Satellite Observations. Geophysical Research Letters 42, 1984e1991. Zech, M., Mayr, C., Tuthorn, M., Leiber-Sauheitl, K., Glaser, B., 2014. Oxygen isotope ratios (18O/16O) of hemicellulose-derived sugar biomarkers in plants, soils and sediments as paleoclimate proxy I: insight from a climate chamber experiment. Geochimica et Cosmochimica Acta 126, 614e623. Zemunik, G., Turner, B.L., Lambers, H., Laliberte´, E., 2015. Diversity of plant nutrientacquisition strategies increases during long-term ecosystem development. Nature Plants 1, 15050. Zemunik, G., Turner, B.L., Lambers, H., Laliberte´, E., 2016. Increasing plant species diversity and extreme species turnover accompany declining soil fertility along a long-term chronosequence in a biodiversity hotspot. Journal of Ecology 104, 792e805. Zeri, M., Sa´, L.D.A., Manzi, A.O., et al., 2014. Variability of carbon and water fluxes following climate extremes over a tropical forest in Southwestern Amazonia. PLoS One 9, e88130. Zhang, K., Kimball, J.S., Nemani, R.R., et al., 2015. Vegetation greening and climate change promote multidecadal rises of global land evapotranspiration. Scientific Reports 5, 15956.

FURTHER READING Bloom, A.J., Asensio, J.S.R., Randall, L., Rachmilevitch, S., Cousins, A.B., Carlisle, E.A., 2012. CO2 enrichment inhibits shoot nitrate assimilation in C3 but not C4 plants and slows growth under nitrate in C3 plants. Ecology 93, 355e367. Nelson, G.C., Bennett, E., Berhe, A.A., et al., 2006. Anthropogenic drivers of ecosystem change: an overview. Ecology and Society 11, 29. Polissar, P.J., Abbott, M.B., Wolfe, A.P., Vuille, M., Bezada, M., 2013. Synchronous interhemispheric Holocene climate trends in the tropical Andes. Proceedings of the National Academy of Sciences of the United States of America 110, 14551e14556. Reich, P.B., Tilman, D., Isbell, F., Mueller, K., Hobbie, S., Flynn, D.F.B., Eisenhauer, N., 2012. Impacts of biodiversity loss escalate through time as redundancy fades. Science 336. Silva, L.C.R., Hoffmann, W.A., Rossatto, D.R., Haridasan, M., Franco, A.C., Horwath, W.R., 2013. Can savannas become forests? A coupled analysis of nutrient stocks and fire thresholds in central Brazil. Plant and Soil 373, 829e842. Vitousek, P.M., Aber, J.D., Howarth, R.W., et al., 1997. Human alteration of the global nitrogen cycle: sources and consequences. Ecological Applications 7, 737e750. Wang, Z., Silva, L.C.R., Sun, G., Luo, P., Mou, C., Horwath, W.R., 2015. Quantifying the impact of drought on soil-plant interactions: a seasonal analysis of biotic and abiotic controls of carbon and nutrient dynamics in high-altitudinal grasslands. Plant and Soil 389, 59e71. Wettstein, J.J., Littell, J.S., Wallace, J.M., Gedalof, Z., 2011. Coherent region-, species-, and frequency-dependent local climate signals in Northern Hemisphere tree-ring widths. Journal of Climate 24, 5998e6012.

Chapter | Three

The Potential for Soils to Mitigate Climate Change Through Carbon Sequestration William R. Horwath*, 1, Yakov Kuzyakovx

*Department of Land, Air and Water Resources, University of California, Davis, CA, United States; xDepartment of Soil Science of Temperate Ecosystems, Department of Agricultural Soil Science, University of Go¨ttingen, Go¨ttingen, Germany 1


INTRODUCTION Soils are the foundation of ecosystems, providing the necessary resources to drive net primary productivity. Soil organic carbon (SOC) is one of the most important characteristics of soils that result from the interplay of net primary producers, decomposers, and mineralogy. The stabilization of SOC with reactive minerals, particularly clays and short-range order metal (hydro)oxide minerals, provides a matrix that supports soil structure, increases water holding capacity, and contains nutrients that allows net primary producers such as crops to thrive. The importance of SOC to support food, energy, and fiber production cannot be understated. Reliance of humanities on soil resources has lead in most cases to its degradation, often in the form of SOC loss. In this chapter, we explore the role and practicality of management practices to support SOC sequestration with the aim of improving soil health and productivity (¼ fertility) and to mitigate carbon dioxide (CO2) emissions from land use change and fossil C combustion.

HUMANITIES RELIANCE AND IMPACT ON SOILS Humanity has relied on soils for fiber and food production for thousands of years (see Brevik chapter). The need to increase food production coincides 61 Climate Change Impacts on Soil Processes and Ecosystem Properties. https://doi.org/10.1016/B978-0-444-63865-6.00003-X Copyright © 2018 Elsevier B.V. All rights reserved.

62

The Potential for Soils to Mitigate Climate Change

FIGURE 3.1 Increasing human population, the numbers of persons supported per hectare, and agricultural land area from 1950 to 2050.

with the increase in the human population. The human population continues to increase from about 2.1 billion in 1950 to a potential of over 10 billion by the midcentury, 2050 (Fig. 3.1). The increase in the human population is supported in part due to the continuous innovation in crop yield improvements and soil management and technologic progress: HabereBosch process for N2 fixation from the atmosphere and NH3 synthesis discovered in the beginning of the 20th century, and widely applied for fertilization after the World War II, maintains now for more than half of global population. The “Green Revolution” in the 1960s was the next major turning point in increasing crop yields. In 1950, each hectare of arable land supported about 2.6 people on a global basis. Today, the intensification of agriculture in the tropics and subtropics of China can support 10e15 people on 1 ha. By 2050, each hectare of land is expected to support 6.2 people (on average worldwide). The increase of food production, intensity per hectare, is mostly limited to the availability of suitable arable land, which will only slightly increase over the period of 1960e2050 from 1.4 to 1.7 million hectares (Fig. 3.2A). Currently, agriculture (cropland) represents 1562 million ha of the total 13,013 million ha of the earth’s land surface. Meadows and pastures (grasslands), forests, and other land (wetlands, deserts, rocky areas) occupy 3406, 3952, and 4093 million ha of the total, respectively (Fig. 3.2). The intensity of crop and soil management since the Green Revolution has increased considerably. This is particularly true for developed countries where agronomic intensification started decades earlier than in developing countries. The increase in agronomic management intensity has resulted in SOC losses of up to 50% globally (Bebi and Brar, 2009). Overall, fossil fuel use to support intensification and the long-term conversion of extensive areas of virgin land

Soil Organic Carbon Balance and Management to Sequester Carbon

63

FIGURE 3.2 The increase in productive agriculture land for different countries (A), and the distribution of agricultural land, grasslands (meadows and pastures), forests, and other land (includes wetlands, rock areas, etc.) for different countries (B).

to agriculture (in the 1930s in the United States, and in the 1950s in Soviet Union) has led to agriculture systems being net contributors that have changed the atmospheric composition and are contributing to climate change (Falkowski et al., 2000).

SOIL ORGANIC CARBON BALANCE AND MANAGEMENT TO SEQUESTER CARBON The general consensus on the carbon balance of agricultural land is that SOC loss has dominated in nearly all areas of the world. Factors leading to SOC

64


FIGURE 3.3 The effect of management practices, rising CO2, and improved crop yields on the degradation of soil organic carbon (SOC) and the effect of management practices to support SOC sequestration since 1950 and projected to 2100.

loss include (1) less input of aboveground and belowground residues by agricultural crops compared with natural vegetation, (2) acceleration of SOC decomposition by tillage disturbance that increases microbial oxidation, (3) relocation of the C input by crops to the upper 20e30 cm soil compared with much deeper roots of native plants (with the consequence of faster decomposition and less long-term stabilization in soil, (4) strong erosion leading to loss of surfacerich organic matter (Fig. 3.3). Other factors, e.g., diminished plant diversity by selecting few specific cash crops, application of chemical fertilizers frequently stimulating SOC decomposition, changing of water regime, etc. are also affecting C sequestration, but are of less importance, and lower microbial populations, leading to reduced input of microbial detritus, the most important source of stable soil C. The initial majority loss of SOC was attributed to the “plow effect” where soil disturbance from tillage exposed protected SOC, particularly occluded within aggregates, subsequently increasing substrate availability followed by microbial decomposition. Other factors discussed or documented in the literature include (1) a reduction in crop species diversity, often a monoculture such as grain crops that reduced the complexity of precursors contributing to stable SOC; (2) a reduction in manure return as farmers specialized in grain crops and eliminated or reduced herd sizes; (3) autumn and winter fallow as fertilizers reduced need for cover crops resulting in a discontinuous plant inputs; (4) as fallow increased soil erosion followed; (5) inefficient nutrient management, such as excess or fall application of nitrogen fertilizers, resulting in loss of reactive N and consequently increased mineralization of the remaining SOC to obtain N; (5) fertilizer effects on the priming of SOC pools (see Yakov’s Chapter); (6) removal of crop residues for animal feed and fuel, resulting in less C inputs and increased soil erosion, and (6) open field burning as a residue management technique to facilitate planting of the next

Animal Manures Sequester Soil Organic Carbon

65

crop, which completely mineralizes aboveground plant residues and also contributed to soil erosion. All of these and other factors contribute to considerable SOC losses from the plow layer of surface soils (0e30 cm) and in many cases deeper. The importance of SOC to soil productivity has long been recognized (Allison, 1973; Jenny, 1941). Numerous studies and observations have concluded that restoring SOC can increase crop yields, particularly in highly degraded soils such as sub-Saharan Africa (Lal, 2007). Cropping and soil management practices, such as reduced or no till, cover crops, diverse crop rotations, manure inputs, etc., can increase SOC (Qiao et al., 2014; Lal, 2004, 2010). However, in a study on arable crops in Europe, restoring SOC did not necessarily increase crop yields, given sufficient nutrients to support productivity are supplied as fertilizer inputs (Hijbeek et al., 2017). The general consensus among soil scientists is that increasing soil C is an important component of increasing soil productivity and in reversing the positive global warming potential (GWP) of agricultural systems.

ANIMAL MANURES SEQUESTER SOIL ORGANIC CARBON The Rothamsted Experiment, started in 1849, showed the positive effects of animal manure application. The annual application of 35 Mg ha1 of animal manure for 140 years showed steady increase in SOC (Jenkinson, 1991). However, the use of animal manure can increase nitrous oxide emissions, but the increase in SOC can often mitigate emissions (Qiao et al., 2014). Numerous studies have shown the positive effects of animal manure on SOC sequestration. This is particularly evident in China, where decadal long-term experiments show the use of animal manure increases SOC and crop yields (Pan et al., 2009). Poudel et al. (2001a) showed that in complex crop rotation systems, the use of composted poultry manure annually significantly increased SOC. The increase in SOC from animal manure application and its effect on crop productivity lasts for decades. However, manure production may deplete soil C from adjacent lands through crop residue removal. In the Rothamsted Experiment, animal manure application for 20 years from 1852 to 1871 resulted in a gradual decline in SOC compared with where no manure was added in the following decades, and after 120 years, SOC remained higher than control pots that never received animal manure (Jenkinson, 1991). Strong positive effects of animal manure on SOC sequestration have been demonstrated universally, provided that it is applied frequently and on a long-term basis. As developing countries increase animal production to satisfy emerging changes in diet, manure production will likely continue to increase and be available as an agricultural soil amendment. However, animal manure use as a soil amendment will be constrained as the agricultural intensification continues. On average, the availability of animal manure for crop production satisfies only 10% of crop N needs globally (Conant et al., 2013) and

66


even less for the stable maintenance of SOC. Therefore, mineral fertilizer N will continue to be used for crop production. Human waste in the form of biosolids may also have the same positive effect on SOC sequestration than does animal manure. Silva et al. (2015) showed that the one time application of 100 Mg ha1 of iron-stabilized biosolids to open cast mine land in Brazil where topsoil was removed exposing the C horizon resulted in SOC sequestration 3e4 times higher than the original savannah. The authors note that after 15 years, new SOC represented 70% of the total SOC with much of C accumulation in small aggregate fractions. The use of biosolids has similar effects to animal manure in promoting SOC sequestration and should be considered as an important nutrient source to enhance soil productivity for crop production in the future.

Conservation Tillage Impact on Soil Organic Carbon Sequestration Conservation tillage incorporates a range of residue management and no till or reduced tillage practices. Conservation tillage practices have been shown to promote SOC sequestration, most notably, in the shallow surface soil (Chambers et al., 2016). Many recent reviews show no significant C sequestration by no-till, rather a redistribution of C from the bottom part of the plow layer to the upper part (see discussion below). Another important benefit of conservation tillage in reducing GWP in addition to removing CO2 from the atmosphere by sequestering soil C is reducing fossil fuel use through reduced tractor passes and other equipment needs (Mitchell et al., 2012). Early studies comparing conservation tillage with tilled soils reported that tilled or disturbed soils released large amounts of CO2 compared with no-till soils (Reicosky and Lindstrom, 1993). Many studies on conservation tillage systems show the majority of sequestered SOC occurs in the shallow 0e15 cm, more often the 0e5 cm soil depth compared to tilled soils. The deposition of SOC in the surface soils under conservation tillage is referred to as stratification. The term stratification ratio has been used routinely in studies to differentiate changes in SOC between tillage systems, and it is defined as the ratio of surface soil (for example, 0e5 cm) to subsurface soil (for example, 15e30 cm) (Franzluebbers, 2002). Studies comparing tillage or organic amendment treatments to depths greater than 30 cm often find no difference in SOC sequestration potential (Horwath et al., 2002). Studies in shallow soil depths under conservation tillage often can overestimate SOC if the comparison depths between treatments do not span at minimum the depth of the tillage. Most conservation tillage studies often do not report SOC values below 15 cm in soil depth. In an overview of conservation agriculture practices and associated ecosystem services, Palm et al. (2014) found that of more than 100 studies, about half reported SOC sequestration rates greater in tilled than no-tilled systems. In addition, a metaanalysis found that crop productivity in cool and wet climates could be reduced in no-till systems (Ogle et al., 2012). The lower crop yields

Animal Manures Sequester Soil Organic Carbon

67

in no-till can reduce crop residue inputs leading to lower SOC sequestration potential in cool mesic climates. Olson (2013) argues to demonstrate unequivocally that SOC sequestration has occurred, equivalent soil mass sampling is needed to assess the impact of tillage (Sollins and Gregg, 2017 Geoderma). For example, a soil core to 1 m in depth is likely needed to determine total SOC change as a result of tillage to verify sole SOC sequestration potential. Another issue associated with the adoption of no till is the initial increase in nitrous oxide emissions due to soil compaction (Six et al., 2004). Often greater than 10 years of no till management is required to mitigate effects of no-till on compaction. Powlson et al. (2014) concluded that no-till systems have limited potential for climate change mitigation, primarily due to limited SOC sequestration potential. However, they also concluded that the accumulation of SOC at the soil surface in no-till system improves soil properties, such as water infiltration and crop growthdboth depending on climate. The stratification of SOC at the surface in no-till systems is a positive ecosystem service that likely overshadows differences in soil C sequestration potential between tillage treatments.

Cover Crops and Crop Rotation Effects on Soil Organic Carbon In many ways, planting cover crops and crop rotations containing grasses have similar outcomes, most notably, sequestering SOC (Poudel et al., 2001). One notable difference in the use of cover crops is that in many regions they provide soil surface cover during fallow periods, such as during winter fallow can reduce erosion and introduce additional soil carbon above and belowground. Similar to a diverse crop rotation, cover crops introduce a diversity of carbon inputs, rooting habits, and rhizodeposits that increase SOC sequestration. Overall, the use of cover crops generally increases SOC and they provide valuable ecosystem services, such as increase water infiltration (Mailapalli et al., 2012), increase nutrient retention, reduce the need for chemical nitrogen inputs, suppress weeds, and increase soil microbial biomass (Drinkwater et al., 1998). Legume cover crops that can biologically fix up to 300 kg N ha1 y1 with rates between 75 and 150 kg N ha1 y1 are more commonly observed (Sarrantonio, 1994; Herridge et al., 2008). For these reasons, cover crops are an essential soil management practice to sustain nutrient availability by increasing SOC and retaining and increasing the availability of nitrogen. Cover crops often increase N use efficiency. In a long-term study, a legume/ grass cover crop in a 4-year crop rotation resulted in accounting for greater than 97% of the biologically fixed and chemical nitrogen inputs over a 10-year period (Poudel et al., 2001). In contrast, in the same study, treatments without cover crops, the applied N recovery was only 75%, but still higher than other systems likely due to a diverse crop rotation. The study demonstrates the value of cover crops and crop rotation in significantly increasing N use efficiency. In other studies, increasing the number of crops in rotation almost universally increases system N use efficiency and SOC sequestration (Liebig et al., 2002;

68


West and Post, 2002). Cover crops can significantly increase available N mineralization for crop uptake and increase microbial biomass and occluded SOC within soil aggregates (Poudel et al., 2001; Sainju et al., 2003; Drinkwater et al., 1998). However, the adoption of cover crops remains low for many reasons. For many farmers, the value of cover crops is well appreciated; however, economic and resource constraints, such as the availability of labor during winter months, establishment issues related to soil moisture and rainfall, and soil management issues related to crop residue management, all factor in decisions farmers are confronted with. Of all the factors, economic outcomes are often the most important consideration for farmers on deciding to plant or not plant a cover crop (Schipanski et al., 2014). Future conservation agricultural systems will likely use cover crops and diverse crop rotations to increase system N use efficiency and SOC as their value in promoting yield stability is appreciated. As mentioned, the use of cover crops essentially increases crop rotation diversity providing multiple benefits that lead to efficient nutrient retention, maintenance or increases in crop yields and minimizes environmental impacts by promoting positive ecosystem services.

POTENTIAL TO SEQUESTER SOIL ORGANIC CARBON The importance of SOC for increasing crop production is well established (Allison, 1973). Interest in using soils to sequester SOC through management with the goal of offsetting anthropogenic CO2 emissions started in the 1990s (Barnwell et al., 1992). As a result, the potential for soil management to sequester SOC has been since extensively studied over the last three decades. The recent 4 per 1000 (4PT) initiative, promoted by the United Nations General Assembly at the 21st Conference of Parties (COP21) held in Paris in 2015, is an integrated effort to promote SOC sequestration as a solution to offset fossil CO2 emissions (see discussion later in chapter). Tables 3.1e3.4 present representative values for SOC sequestration potentials in agriculture, grassland, forest, and wetland systems. The 4PT initiative contends that a small increase in soil C in agriculture, grasslands, and forests would increase crop and fiber production while contributing to offsetting global carbon emissions.

Cropland Agriculture soil represents an important opportunity to sequester SOC. The opportunity arises from the effects of agricultural intensification, which has induced SOC losses through increased mineralization activity and erosion, both water and wind. Agricultural soils are unique in the level of disturbance imparted upon them as a result of inputs and soil management practices. As a result, degraded (but not polluted) agricultural soils have a larger potential sequester SOC. The average lower limit and upper limit for agricultural SOC sequestration is 0.14 and 0.38 Mg g C ha1 y1, respectively (Table 3.1). These potential SOC sequestration rates bracket the values reported in the most recent studies that evaluate the goals set forth in the 4PT initiative (Chambers et al.,

Potential to Sequester Soil Organic Carbon

69

Table 3.1 Representative Agricultural Soil C Sequestration Rates for Different Management Practices and Countries C Sequestration Rate (t C ha1 y1) Low High Reference

Country

Management Practice

Australia

Residue retention

0.15

Lam et al. (2013)

Australia

Reduced tillage

0.34

Sanderman et al. (2010)

Australia

Crop rotation

0.20


Belgium

Farm yard manure

0.45

Buysse et al. (2013)

China

Organic amendment

0.62

Wang et al. (2010)

China

Organic amendment

0.54

Jin et al. (2008)

China

Organic amendment/chemical fertilizer

0.62

Jin et al. (2008)

China

Residue retention

0.57

0.60

Jin et al. (2008)

Canada

Conventional to no-till

0.05

0.16

VandenBygaart et al. (2008)

Canada

Reduced summer fallow

0.30

VandenBygaart et al. (2008)

France

Conservation tillage

0.10

Metay et al. (2009)

France

Crop rotation

0.16

Arrouays et al. (2012, 2014)

India

RiceeRice þ NPK

0.23

Mandal et al. (2008)

India

RiceeRice NPK þ compost

0.41

Mandal et al. (2008)

India

RiceeWheat þ NPK

0.66

Majumder et al. (2008)

India

RiceeWheat þ NPK þ manure

0.99


India

RiceeWheat þ NPK þ residue return

0.89


India

RiceeWheat þ NPK þ green manure

0.82


India

Chemical fertilizer

0.16

Pathak et al. (2011)

India

Chemical fertilizer þ manure

0.33

Pathak et al. (2011)

Indonesia

Chemical fertilizer þ residue return

0.52

Minasny et al. (2012)

Indonesia

Chemical fertilizer þ residue return

0.47


Nigeria

Manure þ residue return

0.30

FAO (2004)

Nigeria

Residue return

0.24

Raji and Ogunwole (2006)

Russia

Crop rotation þ grass

0.08

Savin et al. (2002)

South Korea

Compost

0.24

Lee et al. (2013)

South Korea

Compost þ chemical fertilizer

0.39

Lee et al. (2013)

0.10

0.03

Continued

70


Table 3.1 Representative Agricultural Soil C Sequestration Rates for Different Management Practices and Countries-cont’d C Sequestration Rate (t C ha1 y1) Low High Reference

Country

Management Practice

South Korea

Chemical fertilizer

0.32


United Kingdom

No-till

0.31

Powlson et al. (2012)

United States

No-till

0.40

Johnson et al. (2005)

United States

No-till þ cover crops

0.45

Franzluebbers (2010)

United States

Crop þ ley rotation

0.50

Dick et al. (1998)

United States

Conservation crop rotation

0.15

0.17

Swan et al. (2015)

United States

Residue and tillage management, no-till

0.15

0.27

Swan et al. (2015)

United States

Strip till

0.07

0.17

Swan et al. (2015)

United States

Contour farming

0.07

0.19

Swan et al. (2015)

United States

Cover crop

0.15

0.22

Swan et al. (2015)

United States

Residue and tillage management, reduced till

0.02

0.15

Swan et al. (2015)

United States

Cover crop

0.23

Poudal et al. (2001)

United States

Organic amendment

0.40

Poudal et al. (2001)

United States

Std till þ cover crop

0.20

Mitchell et al. (2016)

United States

No till þ no cover crop

0.51


United States

No till þ cover crop

0.77


Average

0.14

0.38

2016; Minasny et al., 2017). The reader is referred to Minasny et al. (2017) for additional regional potential SOC sequestration rates. There are reasons to suspect that reported potential SOC sequestration values reflect the potential C sequestration under optimal conditions and actually have been overestimated compared with the real agricultural practice. Most reported C sequestration rates were measured on research facilities where management practices are often done year in and year out consistently. For example, the same crop rotation or other management practices are implemented for decades without consideration of economic constraints that often confront farmers. For example, a decadal project performed at the University of California Davis from 1988 to 2000 on the transition from conventional to low-input/ organic agricultural practices determined that low-input system (a hybrid


71

Table 3.2 Representative Grassland Soil C Sequestration Rates for Different Management Practices and Countries

Country

Management Practice

C Sequestration Rate (t C ha1 y1) Low High

Reference

Australia

Crop to pasture

0.30

0.60


Australia

Crop to pasture

0.22

0.76

Chan et al. (2011)

Australia

Crop to pasture

0.78

Badgery et al. (2014)

Australia

Pasture

0.13

Lam et al. (2013)

Australia

Pasture

0.76

Chan et al. (2011)

England

Crop to pasture

0.51

Goulding and Poulton (2005)

France

Crop to pasture

0.49

Arrouays et al. (2002a,b)

France

Permanent pasture

0.10

0.50


France

Improved pasture

0.30

0.40


France

Forage

0.60

0.80

Loiseau and Soussana (1999)

Ireland

Reseeded þ fertilizer N

1.04

1.45

Watson et al. (2007)

Netherlands

Forage

2.29

Loiseau and Soussana (1999)

Switzerland

Forage

0.52

Nitschelm et al. (1997)

United Kingdom

Grassland

1.20

Fitter et al. (1997)

United Kingdom

Grassland on peat soil

6.40

Fitter et al. (1997)

United States

Forage and biomass planting

0.02

0.17

Swan et al. (2015)

United States

Prescribed grazing

0.17

0.44

Swan et al. (2015)

United States

Range planting

0.22

0.35

Swan et al. (2015)

United States

Crop to pasture

0.84

Franzluebbers (2010)

United States

Improved grazing

0.41

Conant et al. (2003)

United States

Grassland

0.45

Potter et al. (1999)

Average

0.33

0.96

between conventional and organic agriculture systems) increased SOC with annual winter cover crops and reduced chemical inputs (fertilizers and pesticides) (Poudel et al., 2001a,b). The low-input management increased system N use efficiency and produced the highest crop yields; however, it was not economically viable compared with the conventional system with no cover crops and exclusive use of fertilizer and chemical inputs. This example demonstrates that researchers can develop appropriate solutions toward climate friendly agricultural systems, but their practical application is constrained by

72


Table 3.3 Representative Forest Soil C Sequestration Rates for Different Management Practices and Countries

Country

C Sequestration Rate Management (t C ha1 y1) Practice Low High

Denmark

Forest

Denmark

0.53

Vesterdal and Raulund-Rasmussen (1998)

Norway spruce

0.35

Vesterdal et al. (2007)

Denmark

Oak

0.08


England

Afforestation

0.57

Poulton et al. (2003)

England

Afforestation

0.54

Poulton et al. (2003)

Finland

Spruce þ pine

0.05

Peltoniemi et al. (2004)

France

Afforestation

0.44


Indonesia

Palm plantation

0.42

Khasanah et al. (2015)

Ireland

Afforestation

2.5

Black et al. (2009)

Nigeria

Afforestation

0.57

Raji and Ogunwole (2006)

Sweden

Forest

0.08

0.28

Berg et al. (2001), Wardle et al. (1997)

Sweden

Scots pine

0.06

0.28

Berg et al. (2001), Staaf and Berg (1997)

Sweden

Norway spruce

0.05

0.65


Taiwan

Afforestation

0.34

Lin et al. (2011)

United States

Afforestation

0.35

Morris et al. (2007)

United States

Afforestation

0.26

Morris et al. (2007)

United States

Forest

0.11

0.43

Kimble et al. (2002)

0.09

0.51

Average

0.14

Reference

not addressing the farmer’s economic situation. A farmer is forced to make decisions based on market pressures and available resources and therefore is unlikely to implement consistent/indefinite management unless it was profitable. Therefore, reported SOC sequestration values provide a good potential but are not guarantees of outcomes at the farm level. The cost of implementing climatefriendly agricultural practices means the need for higher food prices or subsidies to eliminate economic constraints on farmers. Consequently, some state programs, supporting farmers’ decisions focused not only on economic benefits but also on ecosystem functions (e.g., C sequestration), are necessary. Realizing SOC sequestration potential will necessitate consistent (often up to decades) soil or crop management practices. Reducing or eliminating tillage; use of cover crops; use of organic amendments including green compost or


73

Table 3.4 Representative Wetland Soil C Sequestration Rates for Different Types and Countries

Country

Type

C Sequestration Rate (t C ha1 y1) Low High Reference

Australia

Undisturbed

1.05

1.37

Howe et al. (2009)

Australia

Disturbed

0.64

0.89

Howe et al. (2009)

Australia

Mangroves

0.26

3.36

Chmura et al. (2003)

Botswana

Seasonally flooded

0.42

Mitsch et al. (2012)

Canada

Salt marsh

0.63

9.28


Costa Rica

Forest, floodplain

0.84

3.06

Mitsch et al. (2012)

France

Salt marsh

1.61


Finland

Boreal peatland

0.15

0.26

Turunen et al. (2002)

Finland

Temperate peatland

0.1

0.46

Turunen et al. (2002)

Netherlands

Salt marsh

1.39

6.5


North America

Mangroves

1.8


North America

Tidal freshwater

1.4

Craft (2007)

North America

Salt marsh

1.9

Craft (2007)

North America

Everglades

0.86

3.87

Reddy et al. (1993)

North America

Prairie pothole

0.83

3.05

Euliss et al. (2006)

North America

Salt marsh

0.18

7.63


Mexico

Mangrove

1.46

6.54


Southeast Asia

Mangrove

0.9

2.3

Suratman (2008)

Uganda

Cyperus

4.8

Saunders et al. (2007)

United Kingdom

Salt marsh

0.77

1.87


United States, Ohio

Marsh wetland

1.24

2.67

Anderson et al., (2005), Mitsch et al. (2012)

0.75

3.10

Average

manure, crop residue management (i.e., eliminate open field burning or removal), and increasing crop rotation diversity are examples of practices that have been discussed and have been shown to produce positive SOC sequestration results. However, increasing crop rotation diversity can have a negative impact on economic viability if high-value crops are planted less often or displaced by cover crops. This again highlights the economic issues farmers face in implementing practices to promote SOC sequestration.

74


The use of cover crops may not be suitable in all areas due to harsh winter conditions or where cropping system intensity is increased leaving no space in a crop rotation. The availability of organic amendments or their frequency of application maybe limited in areas distant from the facilities that produce them. The elimination of residue management such as open field burning or removal likely has a good potential to increase SOC sequestration, particularly if farmers do not have access to organic amendments or the opportunities to plant cover crops. Of all the above practices, residue management may have the most flexibility in promoting sole SOC sequestration (Blanco-Canqui and Lal, 2009). China provides an excellent example where residues were routinely removed for fodder and fuel. Recent studies show that retention of crop residues in upland soils and rice paddies increased SOC by almost twofold from 3.9e5.7 to 7.1e9.2 g per kilogram of soil over a 15-year period (Li et al., 2010; Wang et al., 2015; Liu et al., 2014). Mandal et al. (2007), showed similar results in India, where they determined the average minimum amount of rice residue retention to maintain current levels of SOC was 2.9 Mg ha1. The increase in crop yields over the last century has resulted mainly from breeding efforts, optimization of NPK fertilization, pesticide use, and rising CO2 levels. The increase in crop yields is increasing crop residue inputs (Pausch and Kuzyakov, 2018 GCB), which may have the potential to increase SOC levels. In India, for example, it was estimated that increases in crop production from nitrogen fertilization, improved crop varieties, and rising CO2 levels increased SOC stocks by 10% (Banger et al., 2015). The need to increase crop biomass to increase yields is a result of stagnating harvest index (HI), defined as the ratio of the weight of grain over total aboveground biomass (grain and aboveground vegetative biomass). The HI of the most intensively cultivated grain crops fall within the range of 0.4e0.6 (Fig. 3.4) (Hay, 1995; Fischer and Edmeades, 2010). Of all the major grain crops, maize has consistently had a higher HI than other cereals. The theoretical ceiling for HI for corn and rice is about 0.54 (McLaughlin et al., 2006;

FIGURE 3.4 Change in harvest index (HI) over time for barley, maize, oats, soybeans, rice, and wheat.


75

Huang et al., 2016), but its linear increase in the last 50 years lets us assume further steady positive development. Soybeans are notable exception to the general trend of increasing HI, showing an inconsistent trend over time. In comparison with other cereal crops, soybeans have an additional carbon sink in roots to support N fixation, which may impact breeding strategies to increase its HI. Because the theoretical HI will likely max out at around 0.55, plant breeding will likely have to increase total biomass to continue to increase yields (McLaughlin et al., 2006). With strategies to increase total biomass to increase grain yields, crop residue production will also increase because of the HI ceiling. Fig. 3.5A shows

FIGURE 3.5 (A) The increase in crop yields over time for barley, maize, oats, soybeans, rice, and wheat. (B) The increase in residue production over time for barley, maize, oats, soybeans, rice, and wheat.

76


the effect of increasing grain yield has had on the comparable increase in crop residue production (Fig. 3.5B) since the beginning of the Green Revolution. The increase in crop residues represents an important source of C inputs that increase SOC. Many studies have shown a direct relationship between increased C inputs and SOC sequestration. The effect of increasing crop residue inputs on SOC levels is highly related to the amount and much less on its type or quality (Larson et al., 1972; Rasmussen and Collins, 1991; Poudel et al., 2001b). On average, less than 10% and often much less of crop residue inputs are converted to SOC (Duiker and Lal, 1999; Kong et al., 2005; Kundu et al., 2007). Legume residues have been reported to increase SOC over nonlegume crop or grass residues (Drinkwater et al., 1998; Poudel et al., 2001b). However, rotations comparing continuous corn versus cornesoybean rotations often show continuous corn with its higher total residue return results in higher SOC (Russell et al., 2005). Singer (2003) found in comparing 125 archive soil samples over a 60-year period in California that intensively managed and irrigated systems have increased in SOC on average from 1.05% to 1.035%, likely reflecting increases in crop productivity and total residue inputs over time. Therefore, total residue inputs maybe more important than the quality of residue inputs in increasing SOC. Increases in crop residue production over time and its effect on SOC are also highly influenced by the accompanying increases in root production and rhizodeposits (Pausch and Kuzyakov, 2018) to support the higher biomass production. Many studies on the effect of crop residues on SOC do not include root residues or rhizodeposits, primarily due to methodological difficulties in assessing belowground inputs (Kuzyakov et al., 2000). Measuring standing root biomass has led to the conclusion that the aboveground represents a larger standing C pool. Rhizodeposits, rarely measured when assessing standing biomass, represent an extremely difficult pool of SOC to assess. Estimates of rhizodeposits range from 2.5 to 6.0 more than standing root biomass (Johnson et al., 2006). Balesdent and Balabane (1996), using 13C enriched corn, estimated rhizodeposition was 1.5 times that of the aboveground biomass. In addition to belowground C inputs, root inputs, although considered a smaller C pool than the aboveground, may also contribute 2e2.5 times more to stable SOC pools (Ka¨tterer et al., 2011; reviewed by Rasse et al., 2005 Geoderma or Plant & Soil). For these reasons, increasing crop residue inputs through strategies to increase total crop biomass may provide a positive interaction through increasing root inputs that are also contributing to increased SOC sequestration potential.

Grasslands Grasslands contain significant SOC stocks. Measured and modeled SOC sequestration rates range from 0 to 8 Mg C ha1 y1 (Jone and Donnelly, 2004). In undisturbed grasslands, total soil C may change little, and the sequestration rates can be viewed as the C inputs required to maintain existing soil C stocks. They occupy about 40.5% of terrestrial land area excluding Greenland


77

and Iceland (World Resource Inst., 2000). Table 3.2 shows typical reported SOC sequestration rates that average 0.33e0.96 Mg C ha1 y1. Belowground SOC storage in grasslands can represent up to 98% of total ecosystem carbon (Hungate et al., 1997; Unteregelsbacher et al., 2012 Biogeochemistry) and nitrogen (Schleuss et al., 2015dEcosystems). The SOC in grasslands can be strongly influenced by management. The transition from cropland to permanent grasslands has the largest potential to sequester SOC (Guo and Gifford, 2002). Introducing productive species that increase carbon inputs can significantly increase SOC sequestration in grasslands (Lal et al., 1998). Practices including fertilization with N and P, irrigation, planned sustainable grazing, and use of legumes, all have the potential to increase SOC sequestration (Conant et al., 2001). However, fertilization with N in grasslands can produce nitrous oxide, offsetting SOC sequestration. Leahy et al. (2004) found over 50% of the CO2 uptake in N-fertilized grasslands was offset by nitrous oxide emissions in Ireland. However, gaseous N losses would likely differ in other soils as a function of climate and soil types. Ruminants, particularly cattle, can emit methane and also offset SOC sequestration in grasslands (O’Mara, 2012). These uncertain offsets can significantly reduce the mitigation offered by increasing SOC sequestration management in grasslands.

Forest Forests contain about 50% of the terrestrial carbon stocks above and belowground (Dixon et al., 1994; Schlessinger, 1997). Established forests have high SOC density. The world area of forests has declined by about 3% since 1990 (Keenan et al., 2015). About 7% of the world forests are intensively managed, planted forests formerly called plantation forests (Birdsey and Pan, 2015). The average SOC sequestration rate in world forest is similar to grasslands at 0.34 Mg C ha1 y1 (Post and Kwon, 2000). As in grasslands, the soil C sequestration in old-growth forests will primarily be used to maintain existing SOC stocks. Table 3.3 shows representative values for SOC sequestration rates ranging from 0.09 to 0.51 Mg C ha1 y1. Afforestation, particularly in degraded or abandoned agricultural land, has the highest SOC sequestration potential. The average C sequestration rate in a broad range of soils of abandoned w45 Mio ha land in Russia amount for 1.0 Mg C ha1 y1 (reviewed Kurganova et al., 2014). Although planted forests have high potential to sequester SOC, they often experience a boom and bust cycle in SOC storage, where short rotation lengths of often less than 10 years leads to loss in SOC on harvest followed by rebuilding a SOC during growth of the next rotation (Lal, 2005). This results in a possible lower potential to accumulate SOC after about 20e30 years (Kurganova et al., 2014). In addition, the planted forests are often fertilized resulting in greenhouse gas (GHG) emissions that can offset SOC sequestration (Sonne, 2006). To avoid GHG emissions that offset SOC sequestration, the N fertilization of planted forests should be avoided.

78


Wetlands Wetlands comprised only 7% of the earth’s terrestrial area but contain some of the highest terrestrial carbon density (Schlesinger, 1997). Destruction wetlands can release enormous amounts of C and nitrous oxide back into the atmosphere. In undisturbed or aggrading wetlands, average SOC sequestration rates can approach 1.18 Mg C ha1 y1 (Mitsch et al., 2012). Table 3.4 shows representative wetlands SOC sequestration rate of 0.75e3.1 Mg C ha1 y1. Therefore, one of the most important practices to avoid release of carbon into the atmosphere is to prevent the destruction and drainage of wetlands. Management options for existing wetlands to increase SOC sequestration potential are limited. One option is to maintain water level in seasonal wetlands to avoid aerobic periods that can lead to SOC oxidation, but this option depends on water availability and surrounding regional practice. The limited availability of highproductive wetland species prevents increasing production potential in wetlands. Restoration of wetlands such as the Everglades and the San Joaquin Sacramento Bay Delta in the United States and other representative tidal areas present opportunities to increase SOC sequestration. However, wetlands can produce significant amounts of methane that may offset SOC sequestration gains (Dalal and Allen, 2008).

SOIL ORGANIC CARBON SEQUESTRATION TO ADDRESS CLIMATE CHANGE Soil organic C is an important component of the global C cycle. Depending on the considered depth, soils contain two to six times the amount of C found in the atmosphere and represent a sink that regulates its CO2 concentration (Houghton, 2007). The decomposition of net primary plant inputs and subsequent production and turnover of microbial biomass produce SOC (Horwath, 2017). The release of stored SOC during land use change and accompanying soil disturbance has released significant amounts of SOC as CO2 back to the atmosphere (Falkowski et al., 2000). In 1850, the atmosphere contained about 285 ppmv CO2 (McCarroll and Loader, 2004) compared with over 400 ppmv or 790 Pg C (also Gt) today. SOC loss to 2 m in depth from the appropriation of land for agricultural use and other land disturbance over the past 12,000 years is estimated to be 133 Pg C (Sandermana et al., 2017). Fossil fuel use, forest burning, and the conversion of extensive areas of virgin land to agriculture are the main sources of terrestrial C released back to the atmosphere (Falkowski et al., 2000). Today, humans are emitting CO2 at a rate of about 9.8e10 Pg C yr1 (Le Que´re´ et al., 2009; Boden et al., 2017). The emission of fossil CO2eC constitutes about 8.5 Gt C yr1 of the total, with the remainder associated with deforestation and land use change. The uptake by oceans has partially offset annual emissions by absorbing 2 Gt in the oceans and 1 Gt of C by plant growth and soil C sequestration (Normile, 2009). Terrestrial SOC sequestration may be underestimated, for example, gallery forests

Sequestering Soil Organic Carbon Requires N

79

expansion into central Brazilian savannas show increases ecosystem C stocks up to 10-fold (Silva et al., 2008). It has been suggested that soils managed to sequester SOC could mitigate some of the effects of fossil C emissions and land use change. Recent agricultural management has been estimated to have released 66 Pg of SOC to the atmosphere (Lal, 1999). At the recent COP21 meeting, the 4PT initiative, “Soils for food security and climate” was embraced. Resequestering this atmospheric CO2 back to SOC could mitigate fossil C emissions to the atmosphere. The estimate of GHG emissions in the 4PT initiative was based on 8.9 Gt of fossil carbon emissions. Global SOC stocks to 2 m in depth are estimated to be 2400 Gt (Batjes, 1996). Therefore the ratio of 8.9 Gt soil C to 2400 global SOC results in the value 0.4% or 4 per mil or 1000. This could limit global warming to an increase of 1.5 C as set forth in the COP21 by capping annual GHG emissions. The worldwide average SOC content per hectare is estimated to be 161 t (Minasny et al., 2017). Assuming that the SOC sequestration rates of the 4PT initiative can be achieved, the average global rate would need to be 0.6 t C ha1 y1. This value falls within the average range of potential SOC sequestration for grasslands, forests, and wetlands but is above the rate for agricultural land. The 4PT initiative value of 0.6 t C ha1 y1 is an average value, suggesting that actual SOC sequestration amounts would be higher or lower depending on soil type and climate. The five soil forming factors would dictate a soil’s potential to sequester SOC depending on the implemented intensity and consistency of management practices.

SEQUESTERING SOIL ORGANIC CARBON REQUIRES N SOC sequestration cannot occur in the absence of N. The C to N ratio of mineral-associated SOC ranges from 8 to 12. There is a direct relationship between SOC and N (Fig. 3.6). Therefore, increasing SOC requires a concomitant increase in soil N. The SOC sequestration rate promoted in the 4PT initiative would require 100 Tg N y1, assuming a soil C to N ratio of 12 (van Groenigen et al., 2017). The authors estimated cropland residue N globally to be about 30 Tg N, well below that needed to form stable SOC. This suggests that additional N is required beyond crop demand to meet the goal of 4PT. The need for N would further be exacerbated if crop residues were removed, for example, for biofuel production (Blanco-Canqui and Lal, 2009). Removing crop residues to produce ethanol would remove N, but also C inputs, reducing SOC sequestration potential or leading to its loss. Fertilizer N inputs have increased over time resulting in increased food production. Despite this increase, estimate of fertilizer N use efficiency in cereals is only 50%. The inefficient use of N in cereal crops suggests that residual N maybe available for SOC sequestration. However, less than 7% of applied fertilizer N is available to subsequent crops, suggesting the N is likely lost from the system via leaching, runoff, and gaseous routes (Ladha et al., 2005). After 40 years of mineral fertilizer N applications that exceeded grain N removal

80

The Potential for Soils to Mitigate Climate Change 8 y = 0.52 + 0.069x R= 0.88 7

Soil N (g kg-1)

6 5 4 3 2 1 0

0

10

20

30 40 50 Soil C (g kg-1)

60

70

80

FIGURE 3.6 The relationship of soil C to N. Data derived from Horwath, W.R., 2017. The role of the soil microbial biomass in cycling nutrients. In: Tate, K.R. (Ed.), Microbial Biomass: A Paradigm Shift in Terrestrial Biogeochemistry. World Scientific, pp. 41e66.

by 60%e190% at the longest continuous maize experiments at the Morrow Plots in Illinois, a net decline in SOC was observed (Khan et al., 2007). In addition, soil N was depleted in the Morrow Plot despite the excess nitrogen inputs (Mulvaney et al., 2009). Monoculture maize could have influenced these results for reasons discussed earlier. Other studies in corn whether mono cropped or in rotation generally show contrasting results with increases in SOC over time (Gregorich et al., 1996; Halvorson et al., 1999; Alvarez, 2005; Clay et al., 2012). The estimated SOC sequestration values shown in Tables 3.2 and 3.3 for grasslands and forests align well with the goals set forth in the 4PT initiative to increase soil C globally on average 0.6 t ha1 y1. Agriculture SOC sequestration alone would only result in about only half expected outcome promoted in the 4PT initiative, and that may be optimistic. However, N availability would also limit SOC sequestration in grasslands and forests. Despite these apparent limitations in the potential to sequester SOC, particularly for the goals set forth in the 4PT initiative, efforts should continue to implement management practices to increase SOC. Increasing SOC as mentioned earlier has positive effects on crop yield potential and is a key in promoting main ecosystem services associated with water and air quality. The management practices in the 4PT initiative promoted our standard conservation agriculture practices that will increase soil productivity through SOC sequestration. These best management practices include afforestation, converted or improved pasture, organic amendments, residue retention, reduced or no-till, crop rotation, and cover crops that have the potential to sequester 0.6, 0.5, 0.5,

Atmospheric Composition and Climate Change Impacts on Soil C

81

0.35, 0.3, 0.2, and 0.35 t C ha1 y1, respectively (Minasny et al., 2017). Overall, the success of the best management practices relies on uniform adoption and consistent year end and year out implementation. Greater success will be seen in degraded soils or soils with lower initial SOC contents. However, there are a number of factors that likely would significantly reduce SOC sequestration potential besides the N limitation described above, most notably, the onset of climate change. Climate change may alter temperature and moisture regimes that could influence decomposition rates and stability of SOC.

ATMOSPHERIC COMPOSITION AND CLIMATE CHANGE IMPACTS ON SOIL C SEQUESTRATION Increasing temperatures and variable magnitude and frequency of precipitation events will have a large impact on crop productivity and microbial processes in the future as well as erosion. The optimistic results of elevated CO2 seen in earlier chamber studies are significantly more compared with free atmospheric carbon exchange experiments (FACE). Ainsworth and Long (2005) found nonsignificant increases in SOC content in FACE wheat and rice experiments. Cotton, a nonfood crop, increased SOC by 42%. Overall, small to no significant changes in SOC have been observed in many FACE studies (Van Kessel et al., 2000). C4 plants such as corn and other grass species showed no response to elevated CO2. A study on global tree growth found that in 60 out of 66 studies, tree growth declined globally despite increasing atmospheric CO2 levels (Silva and Horwath, 2013). Reasons for the decline in tree growth have been debated, but C enrichment of litter may cause N limitation that immobilizes N during decomposition. Elevated CO2 may also compromise the ability of C3 plants including trees to metabolize nitrate, the most widely available form of N in soils. Increased nitrification potential may result from warming of soils following climate change (Butler et al., 2012). Bloom et al. (2012) reported that C3 plants under elevated CO2 have a diminished capacity to reduce nitrate to amines for protein synthesis. Silva et al. (2015) found that coffee trees growing under elevated CO2 also could not metabolize nitrate efficiently. Therefore, the previous consensus that CO2 fertilization derived mainly from small chamber studies would lead to increased plant growth may not be realized for the number of reasons described above. The inability to metabolize nitrate may impact future SOC sequestration capacity if plants cannot achieve full yield potential under elevated CO2. The stability of SOC associated in minerals and occluded within aggregates is also be impacted by a number of factors including increases in temperature, residue quality, and available N. These factors by themselves or through their interactions may impact the decomposition of stable SOC in a process known as the priming effect (PE) (see Yakov’s chapter). The PE was first described by Bingeman et al. (1953), showing that the addition of organic materials influences the mineralization of SOC. Jenkinson (1966) further defined the PE concept as being either positive (increasing SOC mineralization) or negative

82


(decreasing SOC mineralization) in the presence of organic inputs. Overall, the addition of organic materials generally causes a positive priming affect (Jenkinson et al., 1985). The mineralization of CO2 from SOC during the PE also leads to the release of N (Kuzyakov et al., 2000). The addition of crop residues to soils, C-rich root exudates, and other organic materials such as manure cause a positive PE. Some easily decomposable plant materials with low C to N ratios (i.e., living roots), toxic substances (pollutants, pesticides, etc.), and mineral N fertilizers can cause a negative PE. Mechanisms of a PE are complex and may result from the competition for energy and nutrients among specialized microorganisms feeding on simple and complex organic substances (Fontaine et al., 2003). The PE under future climate change is uncertain, but the recent evidence suggests that the temperature sensitivity of SOC will affect stable C pools more intensively as a result of the labile C pool being quickly utilized under warmer soil conditions. However, soil moisture will dictate the magnitude of the response. The majority of SOC is in the stable form, and therefore future increases in temperature may adversely affect soil C sequestration potential (Knorr et al., 2005; Davidson and Janssens, 2006). However, this conclusion is only viable in soils that maintain or increase SOC levels, while disturbed soils would likely decrease in total SOC as has been routinely observed in past studies.

RESEARCH NEEDS IN SOIL ORGANIC CARBON SEQUESTRATION The role of soils in mitigating climate change through SOC sequestration has been studied extensively for the last three decades. Many of the soil management practices discussed here significantly increase SOC sequestration. Most of these results have been obtained from carefully planned studies at research institutions and less so on farmer fields. As suggested, this may overestimate SOC sequestration due to omitting economic constraints that farmers are confronted with as well as not addressing the variability and specifics in management practices and soils resources that affect C sequestration. Grasslands, forests, and wetlands have demonstrated high potential SOC sequestration rates compared with agriculture, not all of these systems can be managed to achieve these potential rates. This is particularly true of systems in equilibrium such as old-growth forests or natural grasslands that would not respond to management practices discussed earlier. Future research on SOC sequestration should include the following: 1. A realistic ecology- and economy-based assessment of the opportunities for farmers to adopt proven management practices to support SOC sequestration. The research should include assessing the barriers to adoption of these practices and to include the most appropriate mechanisms and products needed for farmers to successfully implement appropriate SOC sequestration practices.

References 83 2. The stagnation of the HI in cereal crops implies that increased crop yields will be required to increase grain production. This will result in an increase in crop residue return in future agricultural systems. Although the increased crop residue loads would seemingly benefit SOC sequestration, in many areas of the world this may prove to negatively affect crop productivity by, for example, interfering with planting of the next crop, immobilizing fertilizers, and lowering soil temperatures that impede germination and root development. Therefore, more research is needed on residue management practices that can avoid these issues while simultaneously increasing SOC sequestration. 3. The results from the FACE studies suggest that the elevated CO2 will have little impact on increasing SOC. This is in spite of the fact that small but significant gains in plant productivity have been observed, such as increased rhizodeposition, yet SOC levels are not significantly changed. Therefore, the rate of decomposition of new C inputs has increased suggesting that the previously assumed independence of substrate amount to the decomposition rate constant is no longer a valid assumption and therefore requires additional research. 4. The increased interest in the PE may explain some of the results from FACE studies. A few studies have suggested that stable SOC is not affected by climate change, for example, associated rising temperatures. However, many of these studies are based on laboratory incubations and fail to explain the results observed in FACE studies. Therefore, additional research on the PE is necessary to explain why the increased crop production seen in elevated CO2 experiments and rising temperatures have no effect on SOC sequestration (see Elevated CO2 chapter). 5. Although not discussed in this chapter (see Rodrigues chapter), many of the above research priorities are affected by the soil biology. More research is needed on microbial functions and processes responsible for decomposition of organic inputs and the sequestration and stabilization of SOC. As can be gleaned from the above research priorities, future soil scientists have a variety of research areas to focus on. Soils will play a major role in mitigating climate change; however, the expectations set forth in the 4PT initiative maybe overly optimistic. Despite this optimism, the goal of SOC sequestration is of paramount importance and independent of the expectations of the 4PT. The benefits of promoting SOC sequestration range from maintaining or increasing crop productivity and supporting vital ecosystem services and should be supported through advanced research initiatives.

REFERENCES Ainsworth, E.A., Long, S.P., 2005. What have we learned from 15 years of free-air CO2 enrichment (FACE)? A meta- analytic review of the responses of photosynthesis, canopy properties and plant production to rising CO2. New Phytologist 165, 351e371.

84


Allison, F.E., 1973. Soil Organic Matter and Its Role in Crop Production. In: Developments in Soil Science, vol. 3. Elsevier Scientific, Amsterdam. Alvarez, R., 2005. A review of nitrogen fertilizer and conservation tillage effects on soil organic carbon storage. Soil Use and Management 21, 38e52. https://doi.org/10.1079/ SUM2005291. Anderson, C.J., Mitsch, W.J., Nairn, R.W., 2005. Temporal and spatial development of surface soil conditions in two created riverine marshes. Journal of Environmental Quality 34, 2072e2081. Arrouays, D., Balesdent, J., Germon, J.C., Jayet, P.A., Soussana, J.F., Stengel, P., 2002b. Increasing Carbon Stocks in French Agricultural Soils? Synthesis of an Assessment Report by the French Institute for Agricultural Research on Request of the French Ministry for Ecology and Sustainable Development. Sci. Assess. Unit for Expertise, INRA, Paris (36 pp). Arrouays, D., Grundy, M.G., Hartemink, A.E., Hempel, J.W., Heuvelink, G.B.M., Hong, S.Y., Lagacherie, P., Lelyk, G., McBratney, A.B., McKenzie, N.J., Mendonca-Santos, M.D., Minasny, B., Montanarella, L., Odeh, I.O.A., Sanchez, P.A., Thompson, J.A., Zhang, G.L., 2014. GlobalSoilMap: towards a fine-resolution global grid of soil properties. Advances in Agronomy 125, 93e134. Arrouays, D., Jolivet, C., Boulonne, L., Bodineau, G., Saby, N., Grolleau, E., 2002a. Une initiative nouvelle en France: la mise en place d’un Reseau multi-institutionnel de Mesures de la Qualite des Sols (RMQS). Comptes-rendus des Seances de l’Academie d’Agriculture de France 88 (5), 93e105. Arrouays, D., Marchant, B.P., Saby, N.P.A., Meersmans, J., Orton, T.G., Martin, M.P., Bellamy, P.H., Lark, R.M., Kibblewhite, M., 2012. Generic issues on broad scale soil monitoring schemes: a review. Pedosphere 22, 456e469. Badgery, W.B., Simmons, A.T., Murphy, B.W., Rawson, A., Andersson, K.O., Lonergan, V.E., 2014. The influence of land use and management on soil carbon levels for crop-pasture systems in Central New South Wales, Australia. Agriculture, Ecosystems & Environment 196, 147e157. Balesdent, J., Balabane, M., 1996. Major contribution of roots to soil carbon storage inferred from maize cultivated soils. Soil Biology & Biochemistry 28, 1261e1263. Banger, K., Tian, H., Tao, B., Lu, C., Ren, W., Yang, J., 2015. Magnitude, spatiotemporal patterns, and controls for soil organic carbon stocks in India during 1901e2010. Soil Science Society of America Journal 79, 864e875. https://doi.org/10.2136/sssaj2014.11.0456. Barnwell, T.O., Jackson, R.B., Elliott, E.T., Burke, I.C., Cole, C.V., Paustian, K., Paul, E.A., Donigian, A., Patwardhan, A., Rowell, A., Weinrich, K., 1992. An approach to assessment of management impacts on agricultural soil carbon. Water Air and Soil Pollution 64, 423e435. Batjes, N.H., 1996. Total carbon and nitrogen in the soils of the world. European Journal of Soil Science 47, 151e163. Bebi, D.K., Brar, J.S., 2009. A 25 year record of carbon sequestration and soil properties in intensive agriculture. Agronomy for Sustainable Development 29, 257e265. Berg, B., McClaugherty, C., Virzo De Santo, A., Johnson, D., 2001. Humus buildup in boreal forests e effects of litter fall and its N concentration. Canadian Journal of Forest Research 31, 988e998. Bingeman, C.W., E.Varner, J., Martin, W.P., 1953. The effect of the addition of organic materials on the decomposition of an organic soil. Soil Science Society of America Proceedings 17, 3e38. Birdsey, R., Pan, Y., 2015. Trends in management of the world’s forests and impacts on carbon stocks. Forest Ecology and Management 355, 83e90. Black, K., Byrne, K.A., Mencuccini, M., Tobin, B., Nieuwenhuis, M., Reidy, B., Bolger, T., Saiz, bG., Green, C., Farrell, E.T., Osborne, B., 2009. Carbon stock and stock changes across a Sitka spruce chronosequence on surface-water gley soils. Forestry 82, 255e272. Blanco-Canqui, H., Lal, R., 2009. Crop residue removal impacts on soil productivity and environmental quality. Critical Reviews in Plant Sciences 28, 139e163. https://doi.org/ 10.1080/07352680902776507.

References 85 Bloom, A.J., Asensio, J.S.R., Randall, L., Rachmilevitch, S., CSousins, A.B., Carlisle, E.A., 2012. CO2 enrichment inhibits shoot nitrate assimilation in C3 but not C4 plants and slows growth under nitrate in C3 plants. Ecology 93, 355e367. Boden, T.A., Marland, G., Andres, R.J., 2017. Global, Regional, and National Fossil-Fuel CO2 Emissions. Carbon Dioxide Information Analysis Center, Oak Ridge National Laboratory, U.S. Department of Energy, Oak Ridge, Tenn., U.S.A. https://doi.org/10.3334/ CDIAC/00001_V2017. Butler, S.M., Melillo, J.M., Johnson, J.E., Mohan, J., Steudler, P.A., Lux, H., Burrows, E., Smith, R.M., Vario, C.L., Scott, L., Hill, T.D., Aponte, N., Bowles, F., 2012. Soil warming alters nitrogen cycling in a New England forest: implications for ecosystem function and structure. Oecologia 168, 819e828. https://doi.org/10.1007/s00442-011-2133-7. Buysse, P., Roisin, C., Aubinet, M., 2013. Fifty years of contrasted residue management of an agricultural crop: impacts on the soil carbon budget and on soil heterotrophic respiration. Agriculture, Ecosystems & Environment 167, 52e59. Chambers, A., Lal, R., Paustian, K., 2016. Soil carbon sequestration potential of US croplands and grasslands: implementing the 4 per thousand initiative. Journal of Soil and Water Conservation 71, 68Ae74A. https://doi.org/10.2489/jswc.71.3.68A. Chan, K.Y., Conyers, M.K., Li, G.D., Helyar, K.R., Poile, G., Oates, A., Barchia, I.M., 2011. Soil carbon dynamics under different cropping and pasture management in temperate Australia: results of three long-term experiments. Soil Research 49, 320e328. Chmura, G.L., Anisfeld, S.C., Cahoon, D.R., Lynch, J.C., 2003. Global carbon sequestration in tidal, saline wetland soils. Global Biogeochemical Cycles 17, 1111. https://doi.org/ 10.1029/2002GB001917. Clay, D.E., Chang, J., Clay, S.A., Stone, J., Gelderman, R.H., Carlson, G.C., Reitsma, K., s Jones, M., Janssen, L., Schumacher, T., 2012. Corn yields and No-Tillage affects carbon sequestration and carbon footprints. Agronomy Journal 104, 763e770. Conant, R.T., Berdanier, A.B., Grace, P.R., 2013. Patterns and trends in nitrogen use and nitrogen recovery efficiency in world agriculture. Global Biogeochemical Cycles 27, 558e566. https://doi.org/10.1002/gbc.20053. Conant, R.T., Paustian, K., Elliott, E.T., 2001. Grassland management and conversion into grassland: effects on soil carbon. Ecological Applications 11, 343e355. Conant, R.T., Six, J., Paustian, K., 2003. Land use effects on soil carbon fractions in the southeastern United States. I. Management intensive versus extensive grazing. Biology and Fertility of Soils 38, 386e392. Craft, C., 2007. Freshwater input structures soil properties, vertical accretion, and nutrient accumulation of Georgia and U.S. tidal marshes. Limnology and Oceanography 52, 1220e1230. Dalal, R.C., Allen, D.E., 2008. Greenhouse gas fluxes from natural ecosystems. Australian Journal of Botany 56, 369e407. Davidson, E.A., Janssens, I.A., 2006. Temperature sensitivity of soil carbon decomposition and feedbacks to climate change. Nature 440, 165e173. Dick, W.A., Blevins, R.L., Frye, W.W., Peters, S.E., Christenson, D.R., Pierce, F.J., Vitosh, M.L., 1998. Impacts of agricultural management practices on C sequestration in forest-derived soils of the Eastern Corn Belt. Soil & Tillage Research 47, 235e244. Dixon, R.K., Brown, S., Houghton, R.A., Solomon, A.M., Trexler, M.C., Wisniewski, J., 1994. Carbon pools and flux of global forest ecosystems. Science 263, 185e190. Drinkwater, L.E., Wagoner, P., Sarrantonio, M., 1998. Legume-based cropping systems have reduced carbon and nitrogen losses. Nature 396, 262e265. Duiker, S.W., Lal, R., 1999. Crop residue and tillage effects on carbon sequestration in a Luvisol in central Ohio. Soil & Tillage Research 52, 73e81. Euliss, N.H., Gleason, R.A., Olness, A., McDougal, R.L., Murkin, H.R., Robarts, R.D., Bourbonniere, R.A., Warner, B.G., 2006. North American prairie wetlands are important nonforested land- based carbon storage sites. Science of the Total Environment 361, 179e188.

86


Falkowski, P.R., Scholes, J., Boyle, E., Canadell, J., Caneld, D., Elser, J., Gruber, N., Hibbard, K., Ho¨gberg, P., Linder, S., Mackenzie, F.T., Moore III, B., Pedersen, T., Rosenthal, Y., Seitzinger, S., Smetacek, V., Steffen, W., 2000. The Global Carbon Cycle: a test of our knowledge of Earth as a system. Science 290, 291e296. FAO, 2004. Carbon Sequestration in Dryland Soils. World Soil Resources Reports 102. Food and Agriculture Organization of the United Nations, Rome. Fischer, R.A., Edmeades, G.O., 2010. Breeding and cereal yield progress. Crop Science 50. https://doi.org/10.2135/cropsci2009.10.0564. S-85eS-98. Fitter, A.H., Graves, J.D., Wolfenden, J., Self, G.K., Brown, T.K., Bogie, D., Mansfield, T.A., 1997. Root production and turnover and carbon budgets of two contrasting grasslands under ambient and elevated atmospheric carbon dioxide concentrations. New Phytologist 137, 47e255. Fontaine, S., Mariotti, A., Abbadie, L., 2003. ’The priming effect of organic matter: a question of microbial competition? Soil Biology and Biochemistry 35, 837e843. Franzluebbers, A.J., 2002. Soil organic matter stratification ratio as an indicator of soil quality. Soil & Tillage Research 66, 95e106. Franzluebbers, A.J., 2010. Achieving soil organic carbon sequestration with conservation agricultural systems in the southeastern United States. Soil Science Society of America Journal 74, 347e357. Goulding, K.W.T., Poulton, P.R., 2005. The missing link. Geoscientist 15, 4e7. Gregorich, E.G., Ellert, B.H., Drury, C.F., Liang, B.C., 1996. Fertilization effects on soil organic matter turnover and corn residue C storage. Soil Science Society of America Journal 60, 472e476. Guo, B., Gifford, R.M., 2002. Soil carbon stocks and land use change: a meta analysis. Global Change Biology 8, 345e360. Halvorson, A.D., Reule, C.A., Follett, R.F., 1999. Nitrogen fertilization effects on soil carbon and nitrogen in a dryland cropping system. Soil Science Society of America Journal 63, 912e917. Hay, R.K.M., 1995. Harvest index: a review of its use in plant breeding and crop physiology. Annals of Applied Biology 126, 197e216. Herridge, D.F., Peoples, M.B., Boddey, R.M., 2008. Global inputs of biological nitrogen fixation in agricultural systems. Plant and Soil 311, 1e18. https://doi.org/10.1007/s11104008-9668-3. Hijbeek, R., van Ittersum, M.K., ten Berge, H.F.M., Gort, G., Spiegel, H., Whitmore, A.P., 2017. Do organic inputs matter e a meta-analysis of additional yield effects for arable crops in Europe. Plant and Soil 411, 293e303. Horwath, W.R., Devevre, O.C., Doane, T.A., Kramer, A.W., van Kessel, C., 2002. Soil C sequestration management effects on N cycling and availability (Chapter 14). In: Kimble, J.M., Lal, R., Follett, R.F. (Eds.), Agricultural Practices and Policies for Carbon Sequestration in Soil. Lewis Publishers, pp. 155e164. Horwath, W.R., 2017. The role of the soil microbial biomass in cycling nutrients. In: Tate, K.R. (Ed.), Microbial Biomass: A Paradigm Shift in Terrestrial Biogeochemistry. World Scientific, pp. 41e66. Houghton, R.R., 2007. Balancing the global carbon budget. Annual Review of Earth and Planetary Sciences 35, 313e347. Howe, A.J., Rodriguez, J.F., Saco, P.M., 2009. Surface evolution and carbon sequestration in disturbed and undisturbed wetland soils of the Hunter Estuary, southeast Australia. Estuarine Coastal and Shelf Science 84, 75e83. Huang, M., Chen, J., Cao, F., Jiang, L., Zou, Y., Deng, G., 2016. Improving physiological N-use efficiency by increasing harvest index in rice: a case in super- hybrid cultivar Guiliangyou 2. Archives of Agronomy and Soil Science 62, 725e743. https://doi.org/ 10.1080/03650340.2015.1082031. Hungate, B.A., Holland, E.A., Jackson, R.B., Chapin, F.S., Mooney, H.A., Field, C.B., 1997. The fate of carbon in grasslands under carbon dioxide enrichment. Nature 388, 576e579. Jenkinson, D.S., Fox, R.H., Rayner, J.H., 1985. Interactions between fertilizer nitrogen and soil nitrogendthe so-called ‘priming’ effect. Journal of Soil Science 36, 425e444.

References 87 Jenkinson, D.S., 1991. The Rothamsted long-term experiments: are they still of use? Agronomy Journal 83, 2e10. https://doi.org/10.2134/agronj1991.000 21962008 300010008x. Jenkinson, D.S., 1966. The priming action. In: The Use of Isotopes in Soil Organic Matter Studies. Report of Food and Agriculrure Organisation and International Atomic Energy Agency, Technical Meeting: Braunschweig-VoE lkenrode 1963, pp. 199e207. Jenny, H., 1941. Factors of Soil Formation. McGraw Hill, NewYork, London. Jin, L., Li, Y., Gao, Q., Liu, Y., Wan, Y., Qin, X., Shi, F., 2008. Estimate of carbon sequestration under cropland management in China. Scientia Agricultura Sinica 41, 734e743. Johnson, J.M.-F., Allmaras, R.R., Reicosky, D.C., 2006. Estimating source carbon from crop residues, roots and rhizo-deposits using the national grain-yield database. Agronomy Journal 98, 622e636. Johnson, J.M.F., Reicosky, D.C., Allmaras, R.R., Sauer, T.J., Venterea, R.T., Dell, C.J., 2005. Greenhouse gas contributions and mitigation potential of agriculture in the central USA. Soil & Tillage Research 83, 73e94. Jone, M.B., Donnelly, A., 2004. Carbon sequestration in temperate grassland ecosystems and the influence of management, climate and elevated CO2. New Phytologist 164, 423e439. Ka¨tterer, T., Bolinder, M.A., Andren, O., Kirchmann, H., Menichetti, L., 2011. Roots contribute more to refractory soil organic matter than above-ground crop residues, as revealed by a long-term field experiment. Agriculture, Ecosystems & Environment 141, 184e192. Keenan, R.J., Reams, G.A., Achard, F., de Freitas, J.V., Grainger, A., Lindquist, E., 2015. Dynamics of global forest area: results from the FAO global forest resources assessment 2015. Forest Ecology and Management 352, 9e20. Khan, S.A., Mulvaney, R.L., Ellsworth, T.R., Boast, C.W., 2007. The myth of nitrogen fertilization for soil carbon sequestration. Journal of Environmental Quality 36, 1821e1832. Khasanah, N.M., van Noordwijk, M., Ningsih, H., Rahayu, S., 2015. Carbon neutral? No change in mineral soil carbon stock under oil palm plantations derived from forest or non-forest in Indonesia. Agriculture, Ecosystems & Environment 211, 195e206. Kimble, J., Birdsey, R., Heath, L., Lal, R. (Eds.), 2002. The Potential of US Forest Soils to Sequester Carbon and Mitigate the Greenhouse Effect. CRC Press, Boca Raton, FL. Knorr, W., Prentice, I.C., House, J.I., Holland, E.A., 2005. Long-term sensitivity of soil carbon turnover to warming. Nature 433, 298e301. Kong, A.Y.Y., Six, J., Bryant, D.C., Denison, R.F., van Kessel, C., 2005. The relationship between carbon input, aggregation, and soil organic carbon stabilization in sustainable cropping systems. Soil Science Society of America Journal 69, 1078e1085. https://doi.org/ 10.2136/sssaj2004.0215. Kurganova, V., de Gerenyu, L., Six, J., Kuzyakov, Y., 2014. Carbon cost of collective farming collapse in Russia. Globalk Change Biology 20, 938e947. Kundu, S., Bhattacharyya, R., Prakash, V., Ghosh, B.N., Gupta, H.S., 2007. Carbon sequestration and relationship between carbon addition and storage under rainfed soybeane wheat rotation in a sandy loam soil of the Indian Himalayas. Soil & Tillage Research 92, 87e95. Kuzyakov, Y., Friedel, J.K., Stahr, K., 2000. Review of mechanisms and quantification of priming effects. Soil Biology & Biochemistry 32, 1485e1498. Ladha, J.K., Pathak, H., Krupnik, T.J., Six, J., van Kessel, C., 2005. Efficiency of fertilizer nitrogen in cereal production: retrospect and prospects. Advances in Agronomy 87, 85e156. https://doi.org/10.1016/S0065-2113(05)87003-8. Lal, R., Henderlong, P., Flowers, M., 1998. Forages and row cropping effects on soil organic carbon and nitrogen contents. In: Stewart, B.A. (Ed.), Management of Carbon Sequestration in Soil. CRC Press, Boca Raton, FL, USA, pp. 365e379. Lal, R., 1999. Soil management and restoration for C sequestration to mitigate the greenhouse effect. Progress in Environmental Science 1, 307e326. Lal, R., 2004. Soil carbon sequestration impacts on global climate change and food security. Science 304, 1623e1627.

88


Lal, R., 2005. Forest soils and carbon sequestration. Forest Ecology and Management 220, 242e258. Lal, R., 2007. Anthropogenic influences on world soils and implications to global food security. Advances in Agronomy 93, 69e93. Lal, R., 2010. Enhancing eco-efficiency in agro-ecosystems through soil carbon sequestration. Crop Science 50. S-120eS-131. Lam, S.K., Chen, D., Mosier, A.R., Roush, R., 2013. The potential for carbon sequestration in Australian agricultural soils is technically and economically limited. Scientific Reports 3 (Article number: 2179). Larson, W.E., E.Clapp, C., Pierre, W.H., Morachan, Y.B., 1972. Effects of increasing amounts of organic residues on continuous corn: organic carbon, nitrogen, phosphorus, and sulfur. Agronomy Journal 64, 204e208. Le Que´re´, C., Raupach, M.R., Canadel, J.G., Marland, G., Bopp, L., Ciais, P., Conway, T.J., Doney, S.C., Feely, R.A., Foster, P., Friedlingstein, P., Gurney, K., Houghton, R.A., House, J.I., Huntingford, C., Levy, P.E., Lomas, M.R., Majkut, J., Metzl, N., Ometto, J.P., Peters, G.P., Prentice, I.C., Randerson, J.T., Running, S.W., Sarmiento, J.L., Schuster, U., Sitch, S., Takahashi, T., Viovy, N., van der Werf, G.R., Woodward, F.I., 2009. Trends in the sources and sinks of carbon dioxide. Nature Geoscience 2, 831e836. Leahy, P., Kiely, G., Scanlon, T.M., 2004. Managed grasslands: a greenhouse gas sink or source? Geophysical Research Letters 31, L20507,. https://doi.org/10.1029/ 2004GL021161. Lee, C.H., Jung, K.Y., Kang, S.S., Kim, M.S., Kim, Y.H., Kim, P.J., 2013. Effect of long term fertilization on soil carbon and nitrogen pools in paddy soil. Korean Journal of Soil Science and Fertilization 46, 216e222. Li, Z., Liu, M., Wu, X., Han, F., Zhang, T., 2010. Effects of long-term chemical fertilization and organic amendments on dynamics of soil organic C and total N in paddy soil derived from barren land in subtropical China. Soil & Tillage Research 106, 268e274. Liebig, M.A., Varvel, G.E., Doran, J.W., Wienhold, B.J., 2002. Crop sequence and nitrogen fertilization effects on soil properties in the western corn belt. Soil Science Society of America Journal 66, 596e601. Lin, Y.R., Cheng, C.H., Tseng, T.Y., 2011. Carbon sequestration potential of afforestation in the plain areas in Taiwan: examples from long-term abandoned orchards and afforested sites. Quarterly Journal of Forestry 44 (4), 567e588 (in Chinese). Liu, C., Lu, M., Cui, J., Li, B., Fang, C., 2014. Effects of straw carbon input on carbon dynamics in agricultural soils: a meta-analysis. Global Change Biology 20, 1366e1381. https://doi.org/10.1111/gcb.12517. Loiseau, P., Soussana, J.F., 1999. Elevated CO2, temperature increase and N supply effects on the accumulation of below-ground carbon in a temperate grassland ecosystem. Plant and Soil 212, 123e134. Mailapalli, D.R., Horwath, W.R., Wallender, W.W., Burger, M., 2012. Infiltration, runoff, and export of dissolved organic carbon from furrow-irrigated forage fields under cover crop and No-Till management in the arid climate of California. Journal of Irrigation and Drainage Engineering-ASC 138, 35e42. Majumder, B., Mandal, B., Bandyopadhyay, P.K., Gangopadhyay, A., Mani, P.K., Kundu, A.L., Mazumdar, D., 2008. Organic amendments influence soil organic carbon pools and riceewheat productivity. Soil Science Society of America Journal 72, 775e785. Mandal, B., Majumder, B., Bandyopadhyay, P.K., Hazra, G.C., Gangopadhyay, A., Samantaray, R.N., Mishra, A.K., Chaudhury, J., Saha, M.N., Kundu, S., 2007. The potential of cropping systems and soil amendments for carbon sequestration in soils under long-term experiments in subtropical India. Global Change Biology 13, 357e369. Mandal, B., Majumder, B., Adhya, T.K., Bandyopadhyay, P.K., Gangopadhyay, A., Sarkar, D., Kundu, M.C., Choudhury, S.G., Hazra, G.C., Kundu, S., Samantaray, R.N., Misra, A.K., 2008. Potential of double-cropped rice ecology to conserve organic carbon under subtropical climate. Global Change Biology 14, 2139e2151.

References 89 McCarroll, D., Loader, N.J., 2004. Stable isotopes in tree rings. Quaternary Science Reviews 23, 771e801. McLaughlin, S.B., Kiniry, J.R., Taliaferro, C.M., De La Torre Ugarte, D., 2006. Projecting yield and utilization of swithgrass as an energy crop. Advances in Agronomy 90, 267e297. https://doi.org/10.1016/S0065-2113(06)90007-8. Metay, A., Mary, B., Arrouays, D., Martin, M.P., Nicolardot, B., Germon, J.C., 2009. Effets destechniques culturales sans labour (TCSL) sur le stockage de carbone dans le sol en contexte climatique tempere. Canadian Journal of Soil Science 89 (5), 623e634. Mitchell, J.P., Singh, P.N., Wallender, W.W., Munk, D.S., Wroble, J.F., Horwath, W.R., Hogan, P., Roy, R., Hanson, B.R., 2012. No-tillage and high-residue practices reduce soil water evaporation. California Agriculture 66, 55e61. Minasny, B., Malone, B.P., McBratney, A.B., Angers, D.A., Arrouays, D., Chambers, A., Chaplot, V., Chen, Z.-S., Cheng, K., Das, B.S., Field, D.J., Gimona, A., Hedley, C.B., Hong, S.Y., Mandal, B., Marchant, B.P., Martin, M., McConkey, B.G., Mulder, V.L., O’Rourke, S., Richer-de-Forges, A.C., Odeh, I., Padarian, J., Paustian, K., Pan, G., Poggio, L., Savin, I., Stolbovoy, V., Stockmann, U., Sulaeman, Y., Tsui, C.-C., Va˚gen, T.-G., vanWesemael, B., Winowiecki, L., 2017. Soil carbon 4 per mille. Geoderma 292, 59e86. Minasny, B., McBratney, A.B., Hong, S.Y., Sulaeman, Y., Kim, M.S., Zhang, Y.S., Kim, Y.H., Han, K.H., 2012. Continuous rice cropping has been sequestering carbon in soils in Java and South Korea for the past 30 years. Global Biogeochemical Cycles 26, GB3027. https://doi.org/10.1029/2012GB004406. Mitchell, J.P., Shrestha, A., Mathesius, K., Scow, K.M., Southard, R.J., Haney, R.L., Schmidt, R., Munk, D.S., Horwath, W.R., 2016. Cover cropping and no-tillage improve soil health in an arid irrigated cropping system in California’s San Joaquin Valley, USA. Soil and Tillage Research 165, 325e335. Mitsch, W.J., Bernal, B., Nahlik, A.M., Mander, U., Zhang, L., Anderson, C.J., Jørgensen, S.E., Brix, H., 2012. Wetlands, carbon, and climate change. Landscape Ecology. https://doi.org/10.1007/s10980-012-9758-8. Morris, S.J., Bohm, S., Haile-Mariam, S., Paul, E.A., 2007. Evaluation of carbon accrual in afforested agricultural soils. Global Change Biology 13, 1145e1156. Mulvaney, R.L., Khan, S.A., Ellsworth, T.R., 2009. Synthetic nitrogen fertilizers deplete soil nitrogen: a global dilemma for sustainable cereal production. Journal of Environmental Quality 38, 2295e2314. Nitschelm, J.J., Luscher, A., Hartwig, U.A., van Kessel, C., 1997. Using stable isotopes to determine soil carbon input differences under ambient and elevated atmospheric CO2 conditions. Global Change Biology 3, 411e416. Normile, D., 2009. Round and round: a guide to the carbon cycle. Science 325, 1642e1643. Ogle, M.O., Swan, A., Paustian, K., 2012. No-till management impacts crop productivity. carbon input and soil carbon sequestration 149, 37e49. Olson, K.R., 2013. Soil organic carbon sequestration, storage, retention and loss in U.S. croplands: issues paper for protocol development. Geoderma 195e196, 201e206. O’Mara, F.P., 2012. The role of grasslands in food security and climate change. Annals of Botany 110, 1263e1270. Pan, G., Smith, P., Pan, W., 2009. The role of soil organic matter in maintaining the productivity and yield stability of cereals in China. Agriculture, Ecosystems and Environment 129, 344e348. Pathak, H., Byjesh, K., Chakrabarti, B., Aggarwal, P.K., 2011. Potential and cost of carbon sequestration in Indian agriculture: estimates from long-term field experiments. Field Crops Research 120, 102e111. Palm, C., Blanco-Canqui, H., DeClerck, F., Gatere, L., Grace, P., 2014. Conservation agriculture and ecosystem services: An overview, Agriculture. Ecosystems and Environment 187, 87e105. Pausch, J., Kuzyakov, Y., 2018. Carbon input by roots into the soil: Quantification of rhizodeposition from root to ecosystem scale. Global Change Biology 24, 1e12.

90


Peltoniemi, M., Ma¨kipaä¨, R., Liski, J., Tamminen, P., 2004. Changes in soil carbon with stand age e an evaluation of a modeling method with empirical data. Global Change Biology 10, 2078e2091. Poudel, D.D., Horwath, W.R., Mitchell, J.P., Temple, S.R., 2001. Impacts of cropping systems on soil nitrogen storage and loss. Agricultural Systems 68, 253e268. Post, W.M., Kwon, K.C., 2000. Soil carbon sequestration and land-use change: processes and potential. Global Change Biology 6, 317e328. Potter, K.N., Torbert, H.A., Johnson, H.B., Tischler, C.R., 1999. Carbon storage after longterm grass establishment on degraded soils. Soil Science 164, 718e725. Poudel, D.D., Horwath, W.R., Mitchell, J.P., Temple, S.R., 2001a. Impacts of cropping systems on soil nitrogen storage and loss. Agricultural Systems 68, 253e268. Poudel, D.D., Ferris, H., Klonsky, K., Horwath, W.R., Scow, K.M., van Bruggen, A.H.C., Lanini, W.T., Mitchell, J.P., Temple, S.R., 2001b. The sustainable agriculture farming system project in California’s Sacramento Valley. Outlook on Agriculture 30, 109e116. Poulton, P.R., Pye, E., Hargreaves, P.R., Jenkinson, D.S., 2003. Accumulation of carbon and nitrogen by old arable land reverting to woodland. Global Change Biology 9, 942e955. Powlson, D.S., Stirling, C.M., Jat, M.L., Gerard, B.G., Palm, C.A., Sanchez, P.A., Cassman, K.G., 2014. Limited potential of no-till agriculture for climate change mitigation. Nature Climate Change 4. https://doi.org/10.1038/NCLIMATE2292. Powlson, D.S., Bhogal, A., Chambers, B.J., Coleman, K., Macdonald, A.J., Goulding, K.W.T., Whitmore, A.P., 2012. The potential to increase soil carbon stocks through reduced tillage or organic material additions in England and Wales: a case study. Agriculture, Ecosystems & Environment 146, 23e33. Qiao, Y., Miao, S., Han, X., You, M., Zhu, X., Horwath, W.R., 2014. The effect of fertilizer practices on N balance and global warming potential of maize-soybean-wheat rotations in Northeastern China. Field Crops Research 161, 98e106. Raji, B.A., Ogunwole, J.O., 2006. Potential of soil carbon sequestration under various landuse in the sub-humid and semi-arid savanna of Nigeria: lessons from long-term experiments. International Journal of Soil Science 1 (1), 33e43. Rasmussen, P.E., Collins, H.P., 1991. Long-term impacts of tillage, fertilizer, and crop residue on soil organic matter in temperate semiarid regions. Advances in Agronomy 45, 93e134. Rasse, D.P., Rumpel, C., Dignac, M.-F., 2005. Is soil carbon mostly root carbon? Mechanisms for specific stabilization. Plant and Soil 269, 341e356. Reddy, K.R., DeLaune, R.D., DeBusk, W.F., Koch, M.S., 1993. Long- term nutrient accumulation rates in the Everglades. Soil Science Society of America Journal 57, 1147e1155. Reicosky, D.C., Lindstrom, M.J., 1993. Fall tillage method: effect on short-term carbon dioxide flux from soil. Agronomy Journal 85, 1237e1243. Russell, A.E., Laird, D.A., Parkin, T.B., Mallarino, A.P., 2005. Impact of nitrogen fertilization and cropping system on carbon sequestration in midwestern mollisols. Soil Science Society of America Journal 69, 413e422. Sainju, U.M., Whitehead, W.F., Singh, B.P., 2003. Cover crops and nitrogen fertilization effects on soil aggregation and carbon and nitrogen pools. Canadian Journal of Soil Science 83, 155e165. Sanderman, J., Farquharson, R., Baldock, J., 2010. Soil Carbon Sequestration Potential: A Review for Australian Agriculture. A report prepared for the Department of Climate Change and Energy Efficiency CSIRO National Research Flagships. Sandermana, J., Henglb, T., Fiskea, G.J., 2017. Soil carbon debt of 12,000 years of human land use. PNAS. https://doi.org/10.1073/pnas.1706103114. Sarrantonio, M., 1994. Northeast Cover Crop Handbook. Rodale Institute, Emmaus, PA. Saunders, M.J., Jones, M.B., Kansiime, F., 2007. Carbon and water cycles in tropical papyrus wetlands. Wetlands Ecology and Management 15, 489e498. Savin, I., Sirotenko, O., Romanenkov, V., Shevtsova, L., 2002. Assessment of effect of agrotechnological factors on organic carbon balance in ploughed soils. In: Shishov, L., Voitovich, N. (Eds.), Soils of Moscow Region and Their Usage, pp. 324e335.

References 91 Schipanski, M.E., Barbercheck, M., Douglas, M.R., Finney, D.M., Haider, K., Kaye, J.P., Kemanian, A.R., Mortensen, D.A., Ryan, M.R., Tooker, J., White, C., 2014. A framework for evaluating ecosystem services provided by cover crops in agroecosystems. Agricultural Systems 125, 12e22. Schleuss, P.M., Heitkamp, F., Sun, Y., Miehe, G., Xu, X., Kuzyakov, Y., 2015. Nitrogen Uptake in an Alpine Kobresia Pasture on the Tibetan Plateau: Localization by 15N Labeling and Implications for a Vulnerable Ecosystem. Ecosystems. Schlesinger, W.H., 1997. An Analysis of Global Change. Biogeochemistry. Academic Press, San Diego, CA, USA. Silva, L.C.R., Horwath, W.R., 2013. Explaining global increases in water use efficiency: why have we overestimated responses to rising atmospheric CO2 in natural forest ecosystems? PLoS One 8, e53089. https://doi.org/10.1371/journal.pone.0053089. Silva, L.C.R., Sternberg, L.S.L., Haridasan, M., Hoffman, W.A., Miralles-Wilhelm, F., Franco, A.C., 2008. Expansion of gallery forests into central Brazilian savannas. Global Change Biology 14, 1e11. Silva, L.C.R., Doane, T.A., Corra, R.S., Valverde, V., Pereira, E.I.P., Horwath, W.R., 2015. Iron mediated stabilization of soil carbon amplifies the benefits of ecological restoration in degraded lands. Ecological Applications 25, 1226e1234. Silva, L.C.R., Salamanca-Jimenez, A., Doane, T.A., Horwath, W.R., 2015. Carbon dioxide level and form of soil nitrogen regulate assimilation of atmospheric ammonia in young trees. Scientific Reports 5. https://doi.org/10.1038/srep13141, 13141. Singer, M.J., 2003. Looking back 60 years, California soils maintain overall chemical quality. California Agriculture 57, 38e41. Six, J., Ogle, S.M., Breidt, F.J., Conant, R.T., Mosier, A.R., Paustian, K., 2004. The potential to mitigate global warming with no-tillage management is only realized when practised in the long term. Global Change Biology 10, 155e160. https://doi.org/10.1111/j.15298817.2003.00730.x. Sollins, P., Gregg, J.W., 2017. Soil organic matter accumulation in relation to changing soil volume, mass, and structure: Concepts and calculations 301, 60e71. Sonne, E., 2006. Greenhouse gas emissions from forestry operations: a life cycle assessment. Journal of Environmental Quality 35, 1439e1450. Staaf, H., Berg, B., 1977. A structural and chemical description of litter and humus in the mature Scots pine stand at Ivantja¨rnsheden. Swedish Coniferous Forest Project, Internal Report 65, 31. Suratman, M.H., 2008. Carbon sequestration potential of man-groves in southeast Asia. In: Bravo, F., LeMay, V., Jandl, R., von Gadow, K. (Eds.), Managing Forest Ecosystems. The Challenge of Climate Change. Springer Science, The Netherlands, pp. 297e315. Swan, A., Williams, S.A., Brown, K., Chambers, A., Creque, J.Wick, J., Paustian, K., 2015. Comet-planner. Carbon and Greenhouse Gas Evaluation for NRCS Conservation Practice Planning. A companion report to. www.comet-planner.com. http://comet-planner.nrel. colostate.edu/COMET-Planner_Report_Final.pdf. Turunen, J., Tomppo, E., Tolonen, K., Reinkainen, E., 2002. Estimating carbon accumulation rates of undrained mires in Finland: application to boreal and subarctic regions. The Holocene 12, 79e90. Unteregelsbacher, S., Hafner, S., Guggenberger, G., Siehe, G., Xu, X., Liu, J., Kuzyakoz, Y., 2012. Response of long-, medium- and short-term processes of the carbon budget to overgrazing-induced crusts in the Tibetan Plateau. Biogeochemistry 111, 187e201. van Groenigen, Willem, J., van Kessel, C., Hungate, B.A., Oenema, O., Powlson, D.S., van Groenigen, K.J., 2017. Sequestering soil organic carbon: a nitrogen dilemma. Environmental Science & Technology 51, 4738e4739. € Van Kessel, C., Horwath, W.R., Hartwig, U., Harris, D., LYscher, A., 2000. Net soil carbon input under ambient and elevated CO2 concentrations: isotopic evidence after four years. Global Change Biology 6, 123e135. VandenBygaart, A.J., McConkey, B.G., Angers, D.A., Smith, W., De Gooijer, H., Bentham, M., Martin, T., 2008. Soil carbon change factors for the Canadian agriculture national greenhouse gas inventory. Canadian Journal of Soil Science 88, 671e680.

92


Vesterdal, L., Raulund-Rasmussen, K., 1998. Forest floor chemistry under seven tree species along a soil fertility gradient. Canadian Journal of Forest Research 28, 1636e1647. Vesterdal, L., Ritter, E., Gundersen, P., 2002. Change in organic carbon following afforestation of former arable land. Forest Ecology and Management 169, 137e147. Vesterdal, L., Rosenqvist, L., van der Salm, C., Hansen, K., Groenenberg, B.-J., Johansson, M.-B., 2007. Carbon sequestration in soil and biomass following afforestation: experiences from oak and Norway spruce chronosequences in Denmark, Sweden and The Netherlands. In: Heil, G.W., Muys, B., Hansen, K. (Eds.), Environmental Effects of Afforestation in North-Western Europe e From Field Observations to Decision Support, Plant and Vegetation, vol. 1. Springer, pp. 19e52. Wang, C.J., Pan, G.X., Tian, Y.G., Li, L.Q., Zhang, X.H., Han, X.J., 2010. Changes in cropland topsoil organic carbon with different fertilizations under long-term agro-ecosystem experiments across mainland China. Science China Life Sciences 53, 858e867. Wang, J., Wang, X., Xu, M., Feng, G., Zhang, W., Lu, C., 2015. Crop yield and soil organic matter after long-term straw return to soil in China. Nutrient Cycling in Agroecosystems 102, 371e381. https://doi.org/10.1007/s10705-015-9710-9. Wardle, D.A., Zachrisson, O., Ho¨rnberg, G., Gallet, C., 1997. The influence of island area on ecosystem properties. Science 277, 1296e1299. Watson, C.J., Jordan, C., Kilpatrick, D., McCarney, B., Stewart, R., 2007. Impact of grazed grassland management on total N accumulation in soil receiving different levels of N input. Soil Use and Management 23, 121e128. West, T.O., Post, W.M., 2002. Soil organic carbon sequestration rates by tillage and crop rotation: a global data analysis. Soil Science Society of America Journal 66, 1930e1946. World Resources Institute - PAGE, 2000. Downloaded from: http://earthtrends.wri.org/text/ forests-grasslands-drylands/map-229.htm.

FURTHER READING Blanco-Canqui, H., La, R., 2007. Soil and crop response to harvesting corn residues for biofuel production. Geoderma 141, 355e362. Carter, M.R., 2005. Long-term tillage effects on cool-season soybean in rotation with barley, soil properties and carbon and nitrogen storage for fine sandy loams in the humid climate of Atlantic Canada. Soil and Tillage Research 81, 109e120. Follett, R., Castellanos, J.Z., Buenger, E.D., 2005. Carbon dynamics and sequestration in an irrigated Vertisol in central Mexico. Soil & Tillage Research 83, 148e158. Hunt, P.G., Karlen, D.L., Matheny, T.A., Quisenberry, V.L., 1996. Changes in carbon content of a Norfolk loamy sand after 14 years of conservation or conventional tillage. Journal of Soil and Water Conservation 51, 255e258. Tweeten, L., Thompson, S.R., 2008. Long-Term Agricultural Output Supply-Demand Balance and Real Farm and Food Prices. Working Paper AEDE-WP 0044-08. Ohio State University, Columbus, OH. Yang, J., Zhang, J., 2010. Crop management techniques to enhance harvest index in rice. Journal of Experimental Botany 61, 3177e3189. https://doi.org/10.1093/jxb/erq112.

Chapter | Four

Role of Mineralogy and Climate in the Soil Carbon Cycle

Katherine Heckman*, Craig Rasmussenx, 1

x

*Northern Research Station, USDA Forest Service, Houghton, MI, United States; Soil, Water & Environmental Science Department, University of Arizona, Tucson, AZ, United States 1


MINERALOGY, WEATHERING, AND THE INORGANIC C CYCLE When considering soils in the context of climate regulation, thoughts most often turn to soil organic carbon (SOC) stocks. However, over geologic time scales, a large part of the earth’s C cycle is controlled by the weathering of silicate minerals and carbonates, as well as the formation of pedogenic carbonates. This is the role of mineralogy in the earth’s inorganic C cycle.

Weathering and Consumption of CO2 As parent material is transformed into soil, and primary minerals are dissolved or transformed into secondary minerals, cations are released into the soil solution. This process leads to the conversion of dissolved CO2 to bicarbonate salts which are in turn sequestered over geologic time scales in the form of carbonate rock (Berner et al., 1983; Mackenzie and Lerman, 2006). The contribution of chemical weathering to atmospheric CO2 consumption happens when base cations (Na, K, Ca, Mg) lost during soil development are charge balanced by bicarbonate in soil solution (Chadwick et al., 1994). One mole of base cation charge consumes 1 mol of bicarbonate (e.g., Eq. 4.1). 2ðCO2 Þ þ 3ðH2 OÞ þ CaSiO3 / Ca2þ þ 2 HCO 3 þ H4 SiO4

(4.1) 93

Climate Change Impacts on Soil Processes and Ecosystem Properties. https://doi.org/10.1016/B978-0-444-63865-6.00004-1 Copyright © 2018 Elsevier B.V. All rights reserved.

94


Bicarbonate salts are eventually leached to the ocean where oceanic organisms such as foraminifera, cocolithophorids, corals, and shellfish utilize dissolved bicarbonate salts to form their shells and skeletons, converting these salts to carbonate and carbonic acid. This process releases one of the previously fixed moles of C back into solution (Eq. 4.2). Solid carbonates will eventually be precipitated on the ocean floor and converted to rock. CaðHCO3 Þ2ðaqÞ/CaCO3 ðsÞ þ H2 CO 3 ðaqÞ

(4.2)

Immediate weathering of carbonate parent materials in soils, as opposed to silicate parent materials, both releases and consumes CO2. When Ca and Mg carbonate weather, they release 1 mol of CO2 upon dissolution and consume 2 mol of CO2 when recombining with bicarbonate in soil solution. For example, the C balance of dolomite weathering in soils: 2H2 CO3 ðaqÞ þ Ca; MgðCO3 Þ2ðsÞ/Ca2þ ðaqÞ þ Mg2þ ðaqÞ þ 4HCO 3 ðaqÞ/CaðHCO3 Þ2 þ MgðHCO3 Þ2

(4.3)

The importance of mineral assemblage in determining atmospheric CO2 consumption during pedogenesis is emphasized by several authors (Amiotte-Suchet and Probst, 1993; Bluth and Kump, 1994; Edmond et al., 1995; Garrels and Mackenzie, 1971; Meybeck, 1986), who also highlight the importance of mineral weatherability in determining rates of inorganic C consumption during pedogenesis. Weathering of base-rich extrusive basaltic and ultramafic materials will consume a larger amount of CO2 per mass unit than weathering of more siliceous granitic and andesitic parent materials. For example, across a lithosequence under pine forest, development of soils on basalt parent material consumed 60 kg CO2 m2, in comparison with only 5 kg CO2 m2 consumption during pedogenesis on granite parent material (Heckman and Rasmussen, 2011). Weathering of carbonate materials is a carbon neutral process in the long term, even though it consumes CO2 on shorter time scales (see Eq. 4.2). The inorganic C cycle in soils is also strongly influenced by formation and/or deposition of pedogenic carbonates. High concentrations of pedogenic carbonates commonly occur in soils of arid and semiarid regions. These carbonates may originate from eolian deposition, dissolution of Ca-containing primary minerals, or deposition of Ca-rich rainwater. In contrast to carbonate formation in oceans, pedogenic carbonate formation is mostly an abiotic process, with the reaction being driven by high concentrations of Caþ2, CO3 2 , or HCO3 . Low mean annual precipitation rates, warm temperatures, and dry soil moisture regimes limit water penetration depth and frequency in soils, thereby restricting the dissolution of pedogenic carbonates. These climatic conditions lead to the buildup of significant carbonate deposits over time which can dominate the morphology and function of these soils

Mineralogy, Weathering, and the Inorganic C Cycle

95

(Kraimer et al., 2005; Schlesinger et al., 2008). Carbonate deposits in arid soils also often represent a larger C stock than organic C. For example, in the arid and semiarid soils of Arizona, soil inorganic C storage was nearly three times greater than organic C storage (Rasmussen, 2006).

Inorganic C Stock Sizes and Fluxes The size of inorganic C stocks involved in the weathering and transformation of minerals in soils and sediments can be equivalent or larger than atmospheric and organic C stocks (Table 4.1). The fluxes involved in soil inorganic C cycling are smaller than that of terrestrial organic C cycling but are significant over long time scales. Soil carbonate deposits can have residence times in the order of 10,000 years or more (Schlesinger, 1985) and have shown little sensitivity to shifts in climate, suggesting a stable long-term stock (Monger and Martinez-Rios, 2000). In comparison, SOC stocks cycle more rapidly, typically on a decadal to centurial scale (Amundson, 2001). The Fifth Assessment Report of the IPCC estimates only 0.3 Gt of C is consumed each year due to mineral weathering in soils, in comparison with an estimated flux of 2.6 (1.2) Gt of C sequestered in terrestrial organic reservoirs per year (although the distribution of this sink between aboveground vegetation and soils is not known). In summary, the weathering of primary minerals in parent material and soil acts as an active short-term sink of atmospheric CO2 through formation of aqueous bicarbonates but also acts over geologic time scales to regulate earth’s climate through formation of sedimentary carbonaceous rocks and long-term sequestration of C.

Table 4.1 Carbon Masses in Some Global Reservoirs Reservoir

Gigatons C

Mass of Carbon Moles C

Atmosphere CO2 (preindustrial 280 ppm)

600

5 1016

CO2 (current 400 ppm)

829 10

6.9 1016

Soil organic matter

1920

1.6 1017

Litter and peat

25

2.08 1016

Inorganic soil (CaCO3)

720

6 1016

Ocean CaCO3 surface sediment

65,300,000

5.44 1021

Land

Adapted from Mackenzie, F.T., Lerman, A., 2006. Carbon in the Geobiosphere d Earth’s Outer Shell. Springer, Dordrecht, The Netherlands; with data from IPCC, 2013. Climate change 2013: the Physical Science Basis. In: Stocker, T.F., Qin, D., Plattner, G-K., Tignor, M., Allen, S.K., Boschung, J., Nauels, A., Xia, Y., Bex, V., Midgley, P.M., (Eds.), Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA, 1535 pp.

96


CLIMATE, MINERAL ASSEMBLAGE, AND SOIL ORGANIC CARBON ARE INTRINSICALLY LINKED THROUGH WEATHERING PROCESSES Globally, soils develop and evolve across a continuum of climate regimes and parent materials. The interactions of climate with mineral substrates across time produce a myriad of soil morphologies with contrasting physiochemical environments. These unique suites of physiochemical matrices drive many of the differences observed in SOC storage, the mechanisms controlling C stabilization, and interactions among soil biotic and abiotic components. Pedogenesis and the SOC cycle are both affected by, and linked through, the interactions of climate and mineral substrates.

Climate Versus Physiochemical Control of Soil and Organic C Stock Development The development of mineral soils is driven by energy and mass fluxes constrained by temperature, precipitation, primary production, soil production, and sediment redistribution (Jenny, 1941). The availability of these resources varies widely based on latitude. From a geomorphological perspective, a convenient way to illustrate the consequences of climate for soil profile characteristics is the Strakhov diagram (Fig. 4.1). This diagram summarizes the broad trends in soil depth and mineral character with latitude. In the tropics, where moisture is generally abundant and temperatures are high, chemical weathering rates are at their maximum. The soil weathering zone reaches depths of tens of meters. High precipitation rates produce highly leached soils characterized by high concentrations of Fe and Al oxides. At the opposite side of the spectrum are the taiga and tundra regions of the Arctic. Freezing temperatures limit chemical weathering by keeping water locked up in ice for most of the year. Physical weathering dominates soil development in these regions but

Semi-arid & desert Savannas

Humid tropical

Savannas

Litter production

Evaporation Precipitation

Temperature

Al2O3

Illite-montmorillonite

Fe O3 + Al2O3

Little chemical alteration

Kaolinite

Fresh rock

FIGURE 4.1 Strakhov model (Strakhov, 1967). Reproduced from Rutter (2009) with permission.

Climate, Mineral Assemblage, and Soil Organic Carbon

97

produces little in the way of mineral transformation or morphological development. Here soils are shallow and similar to their parent material in mineral character. In the midlatitudes, factors determining the depth and intensity of the weathering zone are more complex, and relationships between temperature and soil development are less direct. In arid desert regions, temperatures are high, but water limits weathering reactions, and soils remain mostly static in mineral character. In the forests and plains of temperate latitudes, a broad diversity of soil weathering regimes exist which give rise to a high diversity of soil morphology and mineral assemblage. From a pedologic perspective, soil development, here characterized by categorization into one of the 12 soil orders, may be said to follow one of several pathways, or to rephrase, that the degree of soil development is dominated by one of the soil-forming factors: climate, vegetation, topography, parent material, or time (Fig. 4.2). As outlined above, climatic extremes such as tropical, Arctic, and hyperarid regimes give rise to either extremely high or low rates of soil development. In temperate systems, the mineral assemblage of the parent material may be the determining factor in the character of the resulting soil. High amounts of expandable clays in the parent material may result in a vertisol with characteristic shrinking and swelling, resulting in large cracks and constant movement and sheering of large soil aggregates within the soil. Sandy parent materials deposited by glacial retreat in northern temperate zones may develop characteristic spodic horizons due to high water infiltration rates, acidic vegetation, and hydrologic properties of the parent material. SOC stocks, and the stability of those stocks, are dictated by the same soil forming factors that shape soils into their unique morphological and chemical matrices. Interactions among SOC and soil minerals act to decrease the bioavailability of SOC, thereby slowing degradation rates and leading to the preservation of SOC belowground. Extreme climatic conditions lead to extremes in SOC stocks and their turnover. SOC stocks can vary from as little as 300 tons C ha1 in boreal systems where decomposition is limited by soil saturation and low temperatures (Scharlemann et al., 2014). Soil respiration also varies by an order of magnitude or more across the globe, from 80 g C m2 year1 in deserts to 800e2000 g C m2 year1 in a tropical forest (Torn et al., 2009). In desert soils, not only are SOC stocks small, they are generally unstable even when there is an abundance of reactive mineral surfaces. However, large C stocks of inorganic C may be present in these soils and sequestered for long time periods as carbonate minerals. Here, mineralogy (in the form of carbonate minerals) plays a large part in inorganic C deposition and stabilization, but not in organic C stabilization. In arctic and boreal systems, the opposite is true. SOC stocks are generally large and turnover slowly (Fig. 4.3). The preservation of C in these soils is generally mostly dependent on climatic factors, however, not interactions with the minerals, making these C stocks highly vulnerable to future warming.

98


C

Time zero Parent material Wet sites, high water table A O

Site-dominated pathways

C

Cg Cold climate O/A

Entisol

Histosol Climatedominated pathways

Bg

Parent materialdominated pathways

Increasing time and weathering

Cf

Volcanic materials A E

Gelisol

Bh

A Bw

Semi-arid climate Grasslands

C

A C

Btk Bt Mollisol

Andisol

Warm, humid climate Deciduous vegetation A Bt

C

Arid climate Deserts A Bw Bk

Inceptisol

Cool, humid climate Coniferous vegetation E Bs

C

Clayey materials, Wet-dry climate A Btss C

C Vertisol

Alfisol

Spodosol

C Aridisol

Warm, humid climate Older soils A E Bt C Ultisol

Hot, humid climate Very old soils A Bo

Oxisol

FIGURE 4.2 Soil order development pathways. Reproduced with permission from Schaetzl, R.J., Anderson, S., 2005. Soils: Genesis and Geomorphology. Cambridge University Press, New York, 917 pp.

The mineral matrix has received substantial attention in regard to C cycling and stabilization in temperate and tropical regions, where decomposition rates of soil organic matter are thought to be tightly linked to mineralogical parameters. The details and interdependencies of these links between soil C degradation and mineral assemblage are still not fully understood, nor are


99

FIGURE 4.3 Comparison of soil organic carbon (SOC) stock density (kg C per m3 of soil) and SOC mean residence time (MRT) between arid dry ecosystems (desert) and cold moist ecosystems (arctic). These profiles are not meant to be all encompassing but are meant to give some typical examples of the extremes of SOC stocks and stability. Data was taken from Connin, S.L., Virginia, R.A., Chamberlin, C.P., 1997. Carbon isotopes reveal soil organic matter dynamics following arid land shrub expansion. Oecologia 110, 374e386; Lybrand, R.A., Rasmussen, C., 2015. Quantifying climate and landscape position controls on soil development in semiarid ecosystems. Soil Science Society of America Journal 79, 104e116; Lybrand, R.A., Heckman, K., Rasmussen, C., 2018. Soil organic carbon partitioning and residence time variation in desert and conifer ecosystems of southern Arizona. Biogeochemistry (in review); O’Donnell, J.A., Harden, J.W., McGuire, D.A., Kanevskiy, M.Z., Jorgenson, M.T., Xu, X., 2011. The effect of fire and permafrost interactions on soil carbon accumulation in an upland black spruce ecosystem of interior Alaska: implications for post-thaw carbon loss. Global Change Biology 17, 1461e1474; Kaiser, C., Meyer, H., Biasi, C., Rusalimova, O., Barsukov, P., Richter, A., 2007. Conservation of soil organic matter through cryoturbation in arctic soils in Siberia. Journal of Geophysical Research: Biogeosciences 112(G2), doi:10.1029/2006JG000258.

100


the vulnerabilities of these relationships to climate change. However, a comparison among tropical soils can illustrate how the role of the mineral matrix in protecting SOC stocks from biodegradation varies with mineral assemblage within a climate regime. Andisols are among the most organic mattererich mineral soils on the planet, whereas oxisols often have little organic matter that is easily lost after disturbances such as cultivation. The contrast of C density in these tropical soils is one of the most drastic examples of mineral assemblage determining C stock size and stability. Globally, andisols contain nearly three times the C per unit area as oxisols (0.028 vs. 0.010 Pg 1000 km2) (Eswaran et al., 1993). This difference in C content arises from a combination of the unique mineral properties inherited from both parent material and transformation of those minerals by weathering processes. Andisols develop from volcanic parent material containing an abundance of highly weatherable amorphous volcanic glass. The unique chemistry of the parent material, in combination with a wet and warm tropical climate, produces soils enriched in secondary minerals with high specific surface area and charge, capable of binding large amounts of organic material. Tropical oxisols may develop from a number of parent materials, but regardless of the attributes of the parent material, it is time which defines the mineral character of these soils (see Fig. 4.2). Oxisols contain highly weathered materials that have often been eroded and redeposited multiple times prior to morphological development in their current locations, leaving few weatherable mineral phases. Thus the mineral matrix of oxisols is composed mostly of quartz, iron oxides, and aluminum oxides (Van Wambeke et al., 1983; Buol and Eswaran, 1999). Although oxides generally have high specific surface area and pH-dependent charge, pH conditions in the soils lead to low cation exchange capacity (CEC). This low CEC in combination with weathering-induced base cation leaching leads to low fertility and low organic matter contents (Buol et al., 2003). This is only one example of how mineral assemblage can influence SOC stocks within a particular climate regime.

The Chronosequence: Linking SOC Stocks and Mineral Weathering Over Time The current mineral character and C loading of a particular soil is the result of climate acting on parent material over a period of decades to millions of years. Weathering of primary minerals into secondary phases is largely driven by interaction with water and organic acids over time (Huang and Keller, 1972; Drever and Vance, 1994; Dontsova et al., 2014). Organic matter accumulation is concomitant with these weathering processes, itself being driven partially by the creation of new mineral binding sites. Chronosequences can be used to track the coevolution of the mineral matrix and SOC stocks. Perhaps one of the first chronosequences to link C stabilization and the accumulation of reactive secondary mineral phases came from a series of lava flows on the islands of Hawaii, Molokai, and Kauai (Torn et al., 1997). This chronosequence on


101

quickly weathering volcanic substrate tracked the accumulation and eventual loss of reactive noncrystalline mineral phases and SOC stocks over a period of approximately 4 million years. SOC and noncrystalline mineral phases reached a mutual maximum concentration at approximately 150,000 years of soil development, then jointly declined as progressive weathering led to the dissolution and leaching of reactive noncrystalline mineral phases. Similar patterns of accumulation and loss of secondary mineral phases and SOC stocks over geologic time scales can also be seen across chronosequences in temperate climates (Fig. 4.4). The types of secondary mineral or colloidal phases formed over time are necessarily constrained by the parent material, but the importance of Al- and Fe-oxyhydroxide and amorphous phases is a common theme in the long-term accumulation and stabilization of SOC stocks

FIGURE 4.4 Variance in (A) C stocks and (B) stability across a series of chronosequences. Figure legend indicates primary authors and mineral corollary variables.

102


across chronosequences on a range of parent materials. Chronosequences from a grassland over arkosic sedimentary materials (Masiello et al., 2004) and a mixed deciduous forest over loess-mantled glacial outwash (Lawrence et al., 2015) both identified accumulation of organometallic complexes as the strongest determinant of SOC accumulation. A young chronosequence formed through the retreat of the Damma glacier in Switzerland found a close association of SOC stocks and allophane accumulation (Du¨mig et al., 2012). These chronosequences illustrate the links between SOC abundance and reactive mineral abundance across geological time scales.

MINERAL STABILIZATION OF SOIL ORGANIC CdBONDING MECHANISMS Soil minerals vary widely in their elemental composition, degree of crystallinity, and charge characteristics, and therefore in their capacity to form and preserve bonds with organic matter. Most primary minerals are assumed to be fairly unreactive, and therefore not expected to exert a large influence on soil organic matter retention. Secondary mineral phases formed during pedogenesis have received a much larger degree of examination, both in natural soils, artificial soil mixtures, and as pure phases. Secondary phyllosilicates, hydroxides, and short-range order phases are often the focus of most investigations due to their ubiquitous nature in soils and their reactive surface chemistry. Although each secondary mineral has a large range in charge and surface characteristics due to variance in formation conditions among soils, general trends in reactivity and amount of surface area available for sorption are still observable (Table 4.2). Expandable phyllosilicate phases such as smectites have larger surface area and permanent charge than nonexpansible phyllosilicate phases such as kaolinite. Hydroxide phases exhibit higher reactivity than phyllosilicates, and finally short-range order phases, ferrihydrite, allophane, and imogolite, show the highest potential for sorption of organic matter. Organomineral bonding occurs over a spectrum of bond types and strengths. The type and strength of the organomineral bond as well as the organic matter functional groups involved varies based on the mineral phase characteristics outlined above (Mikutta et al., 2007; Wattel-Koekkoek et al., 2001). Bonding mechanisms include (Essington, 2003) (1) Cation exchange, where negatively charged N-containing moieties (quaternized N atom in an aliphatic chain or heterocyclic ring) of the organic matter form a bond with a negatively charged site on the mineral surface; (2) Anion exchange, where protonated mineral surface sites bond with anionic organic groups; (3) Water bridging, the formation of weak H bonds between negatively charged organic functional groups and waters of hydration on a mineral surface; (4) Cation bridging, in which a cation from soil solution acts as a bridge between negatively charged

Mineral Stabilization of Soil Organic CdBonding Mechanisms

103

Table 4.2 Reactive Properties of Common Secondary Soil Minerals Layer Charge (mol per 1/2 Unit Cell) PZNC

SSA (m2 g1)

Cation Exchange Capacity (cmol kg1)

Density

Halloysite

w0

na

21e43

2e60

2.5e2.65

Kaolinite

w0

5.25

5e20

3e15

2.6

Illite

0.6e0.8

3.5e8.5

100e200

10e20

2.6e2.9

Vermiculite (high activity)

0.6e0.9

na

300e500

100e150

2.3e2.7

Smectite

0.2e0.6

na

700e800

80e150

2.0e2.7

Goethite

na

7e9

8e200

0e5

3.3e4.3

Gibbsite

na

9

32 (varies widely)

0e5

2.3e2.4

Ferrihydrite

na

w7.9

100e700

w0

3.8

Allophane

na

5.5e6.9

700e900

10e40

2.3

Imogolite

na

8.4

900e1100

50

2.3

Values taken from Brady and Weil, 2002; Dixon and Schulze, 2002; Cornell and Schwertmann, 2003; Essington, 2003; Joussein et al., 2005; Schaetzl and Anderson, 2005; Karamalidis and Dzombak, 2010; Ma and Karube, 2013; Kleber et al., 2014.

groups on the organic and mineral structures; (5) Ligand exchange, when at low pH, organic matter carboxylate and phenolate groups replace the hydroxyls of protonated mineral surface groups to form strong bonds; (6) Hydrogen bonding, where weak bonds are formed due to the interaction between basal oxygens of mineral surfaces and hydrogens of organic functional groups; (7) van der Waals bonding, when bonds are formed when the fluctuations of the organic and mineral polarizations are correlated; (8) Hydrophobic interaction, when organics precipitate out of soil solution onto mineral surfaces due to their hydrophobic nature. These bonding mechanisms are assumed to be in operation in synthetic and naturally occurring soils, with their relative importance varying with mineral structure, soil solution ionic strength, and pH. Any variance in the susceptibility of these different organomineral complexes to changes in climate has yet to be identified. Outer sphere and electrostatic bonds are assumed to be more vulnerable to disruption, given an increase in temperature, while ligand exchange reactions are thought to be the most stable. Field-scale evidence reflecting changes in SOC stability due to the varying strengths of particular organomineral bonds at the microscale has yet to be presented in a robustly defensible manner.

104


MINERAL STABILIZATION OF SOIL ORGANIC CdFIELD AND LAB-BASED EVIDENCE Data from observational field experiments, lab-based manipulative experiments utilizing field-harvested and artificial soils, as well as batch sorption experiments all suggest a significant influence of the mineral matrix on the binding and preservation of soil organic matter. However, a general agreement among these results has not been reached in regard to the relative importance of the suite of identified mineraleorganic interactions to the preservation of soil organic matter and its interaction with climate.

Lab-Based Studies: Batch Sorption and Incubation Experiments Basic and advanced spectroscopic techniques in combination with batch sorption and incubation studies have allowed for insight into the formation of organomineral bonds. Due to differences in functional groups on the surfaces of minerals and organics, certain organic compounds are hypothesized to bond preferentially to certain mineral phases or certain functional groups on mineral surfaces. As a result of this preferential bonding, when a heterogeneous mixture of organics is reacted with a mineral surface, a phenomenon referred to as “sorptive fractionation” occurs whereby the bulk composition of the mixture of organics is changed after reaction with the mineral. A recent review of the association of organics with mineral particles by Kleber et al. (2014) indicated some trends in the types and strengths of organomineral bonds summarized from lab-based sorption experiments. Fitting with theory, metal oxides formed strong ligand exchange bonds with organic carboxyl groups, whereas weaker bonds such as outer sphere complexation and H bonding were generally observed for reactions of organics with phyllosilicates, although cation bridging was also observed between phyllosilicate surfaces and organic compounds. These results have been achieved through reaction of individual mineral phases with a single organic compound, under controlled conditions. In naturally forming soils, a multitude of complex mineral phases and organic compounds coexist under conditions of fluctuating moisture and organic matter input which leads to swings in pH, organic compound reactivity, soil solution ionic strength, and oxygen availability.

Studies Utilizing Natural Soils: Linkage of Soil Mineralogical and Organic Properties Through Observational Measurements The question remains as to the importance of the above reactions within the context of climatic influence on SOC cycling. The increased retention and stability of soil organic matter due to adsorption on mineral surfaces is well established (e.g., Baldock et al., 2000; Lutzow et al., 2006). However, the relative influence of specific mineral phases in the protection of SOC against

Mineral Stabilization of Soil Organic CdField and Lab-Based Evidence

105

biodegradation is not. There are few studies offering observations of differences in SOC characteristics or bonding as associated with mineral phases in naturally formed soils. Density separation has been employed to separate natural soils into fractions of varying mineral assemblage (Sollins et al., 2006, 2009). This separation scheme utilizes the differences in density among mineral phases (Table 4.2). However, this approach is confounded by the fact that the amount of organic matter associated with the particular mineral fraction decreases the density of the organomineral complexes by an unpredictable degree. Therefore the denser the isolated mineral fraction, the lower the organic matter content. Even considering this complication, separation by density has allowed for the comparison of naturally formed organomineral complexes that differ significantly in their mineral assemblage. Separation of soil fractions dominated by secondary phyllosilicates (less dense) from fractions composed of (assumed to be unreactive) mainly primary minerals (higher density) gave no evidence of mineralogical control over the stability or composition of their associated organics (Sollins et al., 2006). Total C decreased with increasing particle density, but C stability (approximated from radiocarbon content) increased with increasing density. A similar study indicated both increased and decreased stability of C with increasing organomineral complex density (Baisden et al., 2002). In a subsequent investigation by Sollins et al. (2009), again no role of mineral surface chemistry control on SOC stability could be identified in the soils examined, with exception of mineral Fe concentration. Data suggested that Fe oxide phases did protect organic matter from microbial attack for longer periods of time than the other isolated organomineral fractions, evidenced by the relatively unprocessed character and long mean residence time of organics associated with mineral fractions dominated by Fe-dense phases. The question remains, Why does the variance in organomineral bond type and strength observed under laboratory conditions remain so difficult to observe in natural systems? One moderately well-accepted exception to this question occurs in andisols. Andisols and soils with andic properties are perhaps the longest recognized and most studied example of regulation of soil C cycling by mineral and amorphous/sesquioxide phases. Andisols are unique in their high concentrations of allophane and imogolite and poorly crystalline minerals with high reactivity. The high specific surface areas and abundance of reactive surface functional groups of these minerals are thought to be the major driver of SOC accumulation and stabilization in these C dense soils (Harsh et al., 2002; Lilienfein et al., 2004; Rasmussen et al., 2005). Recent publications have used quantification of these poorly crystalline minerals to draw statistical relationships between the abundance of these mineral phases and SOC in andisols (Giardina et al., 2014; Garrido and Matus, 2012). Andisols are a unique case, however, and account for only 3% of soil coverage worldwide (Eswaran et al., 1993). Elucidating the role of mineral surface chemistry in soils of differing morphology remains a challenge.

106


Recent Advances in the Understanding of Mineral Control of Soil Organic Matter at the Landscape Scale On a broad scale, some very general relationships between soil texture, as a proxy for the mineral matrix, and SOC have been in use for decades. Given steady-state inputs of C to the soil, soil texture is generally understood to determine the length of time required for SOC stocks to achieve equilibrium levels (Jenny, 1941). Soil texture is also widely used as an ecosystem scale predictor of SOC stocks in terrestrial biogeochemical models (e.g., CENTURY), with some degree of success. However, significant evidence of mineralogical control of SOC abundance and turnover at the field scale remain difficult to detect due to the influence of confounding environmental factors. New ecosystem- and continental-scale investigations into the importance of soil mineral assemblage in regulation of the SOC cycle are currently under way. Although some trends are emerging, studies are still offering conflicting conclusions. Field studies have begun testing the predictive capacity of soil texture against other soil physiochemical parameters. In a survey of 62 plots on the central US plains, SOC abundance was more strongly related to silt content than clay content. The authors attribute this to the greater water holding capacity of silt-dominated soils which yields better growing conditions and consequently greater C inputs to soils (Augustin and Cihacek, 2016). In a survey of Chilean volcanic soils, texture was not correlated with SOC concentration. Allophane and C complexed with colloidal Al (pyrophosphate extractable Al) explained most of the variation in C content, with the abundance of allophane and CeAl complexes directly dependent on soil pH (Garrido and Matus, 2012). Many times, spatial variability prevents the detection of even general dependencies of C on mineralogical properties. In a recent study of SOC stability and turnover across 13 forests in Switzerland, patterns of turnover (14C abundance) and variance were not related to clay content but exhibited no direct relationship to climatic or geologic variation either. Additionally, in these plots, microtopographical variance in 14C abundance was similar in magnitude to regional-scale variance (van der Voort et al., 2016). The attention of modelers has also turned to the physiochemical characteristics of the soil including mineral assemblage, elemental composition, and chemistry (Bradford et al., 2016). These factors are being introduced as bonding and/or sorption hysteresis reactions which act to modify the cycling rate of soil organic matter pools. The Microbial Efficiency-Matrix Stabilization framework (Cotrufo et al., 2013) suggests that the cycling of microbial decomposition products is explicitly limited through sorption reactions with soil oxides and cation exchange surfaces. No models currently include mineral assemblage explicitly, although efforts toward this end are emerging from exploration of compiled data sets as well as creation of large-scale field observational experiments.

References

107

Recently, examination of top soils (0e10 cm) across a transect of 24 grassland and shrubland sites found significant evidence that adsorption to mineral surfaces drives C accumulation and resistance to biodegradation. Although climatic factors have long been assumed to drive turnover of surface soils, precipitation and temperature were only secondary predictors of C stocks and turnover; geochemical predictors (Si, Al, Fe, concentrations, and base saturation) were more robust predictors (Doetterl et al., 2015). Additionally, some data suggest that mineral assemblage may play more of a role in regulating the stability and turnover of “deep” soil C (>30 cm below the surface). A statistical exploration of 122 soil radiocarbon profiles from all over the world by Mathieu et al. (2015) indicated a strong dependence of soil C turnover on soil morphological characteristics as opposed to climate or land use.

SUMMARY The mineral component of soils is often the largest by mass and volume, with a strong influence over the soil physiochemical conditions under which C preservation and degradation take place. Although it is a well-recognized fact that the mineral matrix of the soil plays a part in the global C cycle, how exactly that part stands to be altered under future climate conditions is not known. Soils contain substantial inorganic and organic C stocks, both of which represent a substantial proportion of the terrestrial C pool. Evidence to date indicates low sensitivity of inorganic C stocks to climate change but has not reached a consistent conclusion in regard to SOC stock vulnerability to changing climate. Modelers and experimentalists alike are currently examining the role of mineral assemblage in regulating the cycling of SOC, but to date no consensus has been reached.

REFERENCES Amiotte-Suchet, P., Probst, J.L., 1993. Flux de CO2 atmosphe´rique consomme´ par alte´ration chimeque continentale. Inluence de la nature de la roche. Comptes Rendus de l’Acade´mie des Sciences, Paris 317, 615e622. Amundson, R., 2001. The carbon budget in soils. Annual Review of Earth and Planetary Sciences 29, 535e562. Augustin, C., Cihacek, L.J., 2016. Relationships between soil carbon and soil texture in the northern Great Plains. Soil Science 181 (8), 1e7. Baisden, W.T., Amundson, R., Brenner, D.L., Cook, A.C., Kendall, C., Harden, J.W., 2002. A multiisotope C and N modeling analysis of soil organic matter turnover and transport as a function of soil depth in a California annual grassland soil chronosequence. Global Biogeochemical Cycles 16 (4), 82.1e82.26. Baldock, J.A., Skjemstad, J.O., 2000. Role of the soil matrix and minerals in protecting natural organic materials against biological attack. Organic Geochemistry 31 (7), 697e710. Berner, R.A., Lasaga, A.C., Garrels, R.M., 1983. The carbonate-silicate geochemical cycle and its effect on the atmospheric carbon-dioxide over the past 100 million years. American Journal of Science 283, 641e683. Bluth, G.J.S., Kump, L.R., 1994. Lithological and climatological controls of river chemistry. Geochimica et Cosmochimica Acta 58, 2341e2359.

108


Bradford, M.A., Wider, W.R., Bonan, G.B., Fierer, N., Raymond, P.A., Crowther, T.W., 2016. Managing uncertainty in soil carbon feedbacks to climate change. Nature Climate Change 6, 751e758. Brady, N.C., Weil, R.R., 2002. The Nature and Properties of Soils. Pearson, Boston, USA, 980 pp. Buol, S.W., Eswaran, H., 1999. Oxisols. Advances in Agronomy 68, 151e195. Buol, S.W., Southard, R.J., Graham, R.C., McDaniel, P.A., 2003. Soil Genesis and Classification. Iowa State Press, Ames, Iowa, 494 pp. Connin, S.L., Virginia, R.A., Chamberlin, C.P., 1997. Carbon isotopes reveal soil organic matter dynamics following arid land shrub expansion. Oecologia 110, 374e386. Cornell, R.M., Schwertmann, U., 2003. The Iron Oxides: Structure, Properties, Reactions, Occurrences and Uses. WILEY-VCH GmbH & Co., KGaA, Weinheim, 703 pp. Cotrufo, M.F., Wallenstein, M.D., Boot, C.M., Denef, K., Paul, E., 2013. The Microbial Efficiency-Matrix Stabilization (MEMS) framework integrates plant litter decomposition with soil organic matter stabilization: do labile plant inputs form stable soil organic matter? Global Change Biology 19 (4), 988e995. Dixon, J.B., Schulze, D.G., 2002. Soil Mineralogy with Environmental Applications. Soil Science Society of America, Inc., Madison, Wisconsin, USA, 866 pp. Doetterl, S., Stevens, A., Six, J., Merckx, R., Van Oost, K., Pinto, M.C., CasanovaKatny, A., Munõz, C., Boudin, M., Venegas, E.Z., Boeckx, P., 2015. Soil carbon storage controlled by interactions between geochemistry and climate. Nature Geoscience 8, 780e783. Drever, J.I., Vance, G.F., 1994. Role of soil organic acids in mineral weathering processes. In: Pittman, E.D. (Ed.), Organic Acids in Geological Processes. Springer-Verlag, Berlin, Heidelberg, pp. 138e161. Du¨mig, A., Haüsler, W., Steffens, M., Ko¨gel-Knabner, I., 2012. Clay fractions from a soil chronosequence after glacier retreat reveal the initial evolution of organoemineral associations. Geochimica et Cosmochimica Acta 85, 1e18. Edmond, J.M., Palmer, M.R., Measures, C.I., Grant, B., Stallard, R.F., 1995. Fluvial geochemistry and denudation rate of Guyana Shield. Geochimica et Cosmochimica Acta 59, 3301e3325. Essington, M.E., 2003. Soil and Water Chemistry: An Integrative Approach. CRC Press, Boca Raton, USA, 534 pp. Eswaran, H., Van Den Berg, E., Reich, P., 1993. Organic carbon in soils of the world. Soil Science Society of America Journal 57, 192e194. Garrels, R.M., Mackenzie, F.T., 1971. Evolution of Sedimentary Rocks. Norton, New York, USA, 397 pp. Garrido, E., Matus, F., 2012. Are organo-mineral complexes and allophane content determinant factors for the carbon level in Chilean volcanic soils? Catena 92, 106e112. Harsh, J., Chorover, J., Nizeyimana, E., 2002. Allophane and imogolite. In: Dixon, J.B., Schulze, D.G. (Eds.), Soil Mineralogy with Environmental Applications. Soil Science Society of America Inc, Madison, Wisconsin, USA, pp. 291e322. Heckman, K., Rasmussen, C., 2011. Lithologic controls on regolith weathering and mass flux in forested ecosystems of the southwestern USA. Geoderma 164, 99e111. Huang, W.H., Keller, W.D., 1972. Organic acids as agents of chemical weathering of silicate minerals. Nature 239 (96), 149e151. IPCC, 2013. Climate change 2013: the physical science basis. In: Stocker, T.F., Qin, D., Plattner, G.-K., Tignor, M., Allen, S.K., Boschung, J., Nauels, A., Xia, Y., Bex, V., Midgley, P.M. (Eds.), Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA, 1535 pp. Kaiser, C., Meyer, H., Biasi, C., Rusalimova, O., Barsukov, P., Richter, A., 2007. Conservation of soil organic matter through cryoturbation in arctic soils in Siberia. Journal of Geophysical Research: Biogeosciences 112 (G2). https://doi.org/10.1029/2006JG000258. Karamalidis, A.K., Dzombak, D.A., 2010. Surface Complexation Modeling: Gibbsite. John Wiley & Sons Inc, Hoboken, New Jersey, 198 pp.

References

109

Kleber, M., Eusterhues, K., Keiluweit, M., Mikutta, C., Mikutta, R., Nico, P.S., 2014. Mineralorganic associations: formation, properties, and relevance in soil environments. Advances in Agronomy 130. ISSN 0065-2113. Kraimer, R.A., Monger, H.C., Steiner, R.L., 2005. Mineralogical distinctions of carbonates in desert soils. Soil Science Society of America Journal 69 (6), 1773e1781. Lawrence, C.R., Harden, J.W., Xu, X., Schulz, M.S., Trumbore, S.E., 2015. Long-term controls on soil organic carbon with depth and time: a case study from the Cowlitz River Chronosequence, WA USA. Geoderma 247, 73e87. Lilienfein, J., Qualls, R.G., Uselman, S.M., Bridgham, S.D., 2004. Adsorption of dissolved organic carbon and nitrogen in soils of a weathering chronosequence. Soil Science Society of America Journal 68, 292e305. Lu¨tzow, M.V., Ko¨gel-Knabner, I., Ekschmitt, K., Matzner, E., Guggenberger, G., Marschner, B., Flessa, H., 2006. Stabilization of organic matter in temperate soils: mechanisms and their relevance under different soil conditionsea review. European Journal of Soil Science 57 (4), 426e445. Lybrand, R.A., Rasmussen, C., 2015. Quantifying climate and landscape position controls on soil development in semiarid ecosystems. Soil Science Society of America Journal 79, 104e116. Lybrand, R.A., Heckman, K., Rasmussen, C., 2018. Soil organic carbon partitioning and residence time variation in desert and conifer ecosystems of southern Arizona. Biogeochemistry (in review). Ma, Y.L., Karube, J., 2013. Imogolite flocculations under alkaline conditions. Soil Science and Plant Nutrition 59 (2), 125e129. Mackenzie, F.T., Lerman, A., 2006. Carbon in the Geobiosphere d Earth’s Outer Shell. Springer, Dordrecht, The Netherlands. Masiello, C.A., Chadwick, O.A., Southon, J., Torn, M.S., Harden, J.W., 2004. Weathering controls on mechanisms of carbon storage in grassland soils. Global Biogeochemical Cycles 18 (4), 1e9. Mathieu, J.A., Hatte, C., Balesdent, J., Parent, E., 2015. Deep soil carbon dynamics are driven more by soil type than by climate: a worldwide meta-analysis of radiocarbon profiles. Global Change Biology 21, 4278e4292. Meybeck, M., 1986. Composition des ruisseau non pollue´s de France. Sciences Geólogiques, bulletins et me´moires 39, 3e77. Mikutta, R., Mikutta, C., Kalbitz, K., Scheel, T., Kaiser, K., Jahn, R., 2007. Biodegradation of forest floor organic matter bound to minerals via different binding mechanisms. Geochimica et Cosmochimica Acta 71, 2569e2590. Monger, H.C., Martinez-Rios, J.J., 2000. Chapter 4: inorganic carbon sequestration in grazing lands. In: Follett, R.F., Kimble, J.M., Lal, R. (Eds.), The Potential of U.S. Grazing Lands to Sequester Carbon and Mitigate the Greenhouse Effect. Lewis Publishers, Boca Raton, FL, pp. 87e118. O’Donnell, J.A., Harden, J.W., McGuire, D.A., Kanevskiy, M.Z., Jorgenson, M.T., Xu, X., 2011. The effect of fire and permafrost interactions on soil carbon accumulation in an upland black spruce ecosystem of interior Alaska: implications for post-thaw carbon loss. Global Change Biology 17, 1461e1474. Rasmussen, C., Torn, M.S., Southard, R.J., 2005. Mineral assemblage and aggregates control carbon dynamics in a California conifer forest. Soil Science Society of America Journal 69 (6), 1711e1721. Rasmussen, C., 2006. Distribution of soil organic and inorganic carbon pools by biome and soil taxa in Arizona. Soil Science Society of America Journal 70, 256e265. Schaetzl, R.J., Anderson, S., 2005. Soils: Genesis and Geomorphology. Cambridge University Press, New York, 917 pp. Scharlemann, J.P.W., Tanner, E.V.J., Hiederer, R., Kapos, V., 2014. Global soil carbon: understanding and managing the largest terrestrial carbon pool. Carbon Management 5 (1), 81e91. Schlesinger, W.H., 1985. The formation of caliche in soils of the Mojave Desert, California. Geochimica et Cosmochimica Acta 49 (1), 57e66.

110


Schlesinger, W.H., Belnap, J., Giles, M., 2008. On carbon sequestration in desert ecosystems. Global Change Biology 15 (6), 1488e1490. Sollins, P., Swanston, C., Kleber, M., Filley, T., Kramer, M., Crow, S., Calwell, B.A., Lajtha, K., Bowden, R., 2006. Organic C and N stabilization in a forest soil: evidence from sequential density fractionation. Soil Biology and Biochemistry 38 (11), 3313e3324. Sollins, P., Kramer, M.G., Swanston, C., Lajtha, K., Filley, T., Aufdenkampe, A.K., Wagai, R., Bowden, R.D., 2009. Sequential density fractionation across soils of contrasting mineralogy: evidence for both microbial-and mineral-controlled soil organic matter stabilization. Biogeochemistry 96 (1e3), 209e231. Torn, M.S., Trumbore, S.E., Chadwick, O.A., Vitousek, P.M., Hendricks, D.M., 1997. Mineral control of soil organic carbon storage and turnover. Nature 389 (6647), 170e173. Torn, M.S., Swanston, C.W., Castanha, C., Trumbore, S.E., 2009. Storage and turnover of natural organic matter in soil. In: Senesi, N., Baoshan, X., Huange, P.M. (Eds.), Biophysico-chemical Processes Involving Natural Nonliving Organic Matter in Environmental Systems. John Wiley & Sons, Inc, pp. 219e272. van der Voort, T.S., Hagedorn, F., McIntyre, C., Zell, C., Walthert, L., Schleppi, P., Feng, X., Eglinton, T.I., 2016. Variability in 14C contents of soil organic matter at the plot and regional scale across climatic and geologic gradients. Biogeosciences 13, 3427e3439. Van Wambeke, A., Eswaran, H., Herbillon, A.J., Comerma, J., 1983. Oxisols. Developments in Soil Science 11, 325e354. Wattel-Koekkoek, E.J.W., van Genuchten, P.P.L., Buurman, P., van Lagen, B., 2001. Amount and composition of clay-associated soil organic matter in a range of kaolinitic and smectitic soils. Geoderma 99, 27e49.

Chapter | Five

Impacts of Climate Change on Soil Microbial Communities and Their Functioning

Franciska T. de Vries*, 1, Robert I. Griffithsx

*School of Earth and Environmental Sciences, The University of Manchester, Manchester, United Kingdom; xCentre for Ecology and Hydrology, Wallingford, United Kingdom 1


INTRODUCTION The importance of soil microbial communities for ecosystem functioning is now firmly established. Yet, despite their importance for processes of carbon (C) and nutrient cycling and structuring plant communities (Van der Heijden et al., 2008), we still cannot reliably predict how soil microbial communities and their functioning will be affected by climate change. Climate change, caused by elevated levels of atmospheric CO2, is resulting in changed precipitation patterns, increased temperatures, reduced snow cover, and an increased frequency and intensity of extreme weather events (IPCC, 2014). All these climate change drivers can affect soil microbial communities either directly, or indirectly through affecting plant physiological processes and plant community composition (Bardgett et al., 2013). And while we broadly understand the mechanisms through which climate change can affect microbial communities, we are still in the dark when it comes to predicting microbial community response to individual and interactive climate change drivers, let alone the consequences for ecosystem functioning. This is an important knowledge gap because the activities of soil microbial communities control, to a large extent, the balance between the storage and release of greenhouse gases such as CO2 and N2O, and can therefore exacerbate or mitigate climate change. Here, we summarize and synthesize what is currently known about soil microbial community response to climate change and propose how we might use microbial Climate Change Impacts on Soil Processes and Ecosystem Properties. https://doi.org/10.1016/B978-0-444-63865-6.00005-3 Copyright © 2018 Elsevier B.V. All rights reserved.

111

112

Impacts of Climate Change on Soil Microbial Communities

growth strategies and functional traits to predict both their response to climate change and the consequences for soil functioning.

A SHORT HISTORY OF RESEARCH ON CLIMATE CHANGE IMPACTS ON SOIL MICROBIAL COMMUNITIES In the beginning of the 1990s, the ecological community started to realize that the dawning and imminent change in climate would affect the future form and functioning of the world’s ecosystems. In their landmark paper, Raich and Schlesinger (1992) predicted that increasing temperatures would increase rates of soil respiration, thereby creating a positive feedback to climate change. At that time, evidence was starting to emerge that soil bacteria and fungi contribute differently to organic matter decomposition and soil respiration, primarily through studies in which either fungal or bacterial growth was inhibited through the addition of antibiotics (Anderson and Domsch, 1973; Nakas and Klein, 1980). While soil scientists also started to find evidence that soil management and soil moisture content could affect the relative proportion of fungi and bacteria in the soil microbial community (Nakas and Klein, 1979; Schnurer et al., 1986), these early studies on how climate change affects soil functioning considered the soil as a “black box.” Soil organisms and their functioning were mostly ignored because methodological constraints prevented a detailed exploration of the functional roles and responses of different biotic components to change. However, the following decades would see an exponential growth in studies assessing the mechanisms through which climate change affects soil functioning, aided by a revolution in the use of both stable isotope methods that allow the tracing of elements through different components of the soil system, and molecular and metagenomic approaches that allow for the identification of changes in abundance, activity, and functional potential of soil microbial communities. Moreover, more and more studies would consider the interactions between plants and soil communities under these changing environmental conditions and the implications for ecosystem functioning.

Elevated CO2 The first studies to assess the consequences of climate change for soil functioning focused on the effects of elevated CO2. Several studies in the 1990s found that plant growth and belowground allocation of C, particularly of rhizodeposits, increased under elevated CO2, and that this had consequences for microbial biomass and respiration rates (Zak et al., 1993; Newton et al., 1995). Zak et al. (1996) were the first to focus on the effects of elevated CO2 on microbial community composition. These authors hypothesized that the proportion of fungi would increase under elevated CO2 because of increased plant litter production. They assessed shifts in microbial community composition using Phospholipid Fatty Acid (PLFA) analysis, which was the first chemotaxonomic method to assess shifts in total fungal and bacterial abundance as well as shifts in bacterial community composition based on lipid biomarkers

A Short History of Research on Climate Change Impacts

113

(Frostega˚rd et al., 1993). However, they found no evidence that elevated CO2 altered soil microbial community composition; a result that they would repeat in a subsequent experiment (Zak et al., 2000). In contrast, in a similar experiment that also involved N addition, Klironomos et al. (1996) found that under low N conditions, elevated CO2 increased both decomposer and AM fungal abundance, assessed using microscopic counts of observed hyphae. In addition, Jones et al. (1998) found that elevated CO2 affected fungal community composition, as analyzed using culturing methods. Specifically, these authors found that cellulose decomposers increased under elevated CO2, and that this resulted in increased decomposition rates. However, they also found that elevated CO2 had no effect on bacterial community composition as analyzed by an early type of DNA-based community profiling. These DNA-based approaches allowed for more detailed investigation of changes in microbial communities in response to elevated CO2, for example, assessing discrete changes within fungal and bacterial communities. Lipson et al. (2006) sequenced 149 bacterial clones and found that alphaproteobacteria, beta-proteobacteria, and Acidobacteria increased, and that variation in bacterial clones was strongly linked to microbial functioning. A year later, Drigo et al. (2007) were the first to combine quantitative Polymerase Chain Reaction (PCR) and PCR-denaturing gradient gel electrophoresis (DGGE) to assess the effects of elevated CO2 on soil microbial communities. These authors found that elevated CO2 significantly affected the size and composition of both bacterial and fungal communities, but that this effect depended strongly on the origin of the soil used and the plant species growing during the experiment. They found that bacterial community composition was affected more strongly than fungal community composition, and attributed this to quantitative and qualitative changes in root exudation. In addition, elevated CO2 increased arbuscular mycorrhizal (AM) fungal abundance (measured as PLFA and NLFA 16:8w5), and bacterial community composition was more strongly affected in mycorrhizal plants, potentially through AM fungi modifying patterns of root exudation. These patterns would be confirmed by later studies that found that elevated CO2 increases total microbial abundance (He et al., 2010; Blankinship et al., 2011) and the relative abundance of fungi (but see Van Groenigen et al., 2007; Cheng et al., 2011; Guenet et al., 2012). While it is clear that elevated CO2 causes changes in fungal and bacterial community composition through affecting the quantity and quality of plant C allocation belowground, the effects on fungal and bacterial diversity are largely inconclusive (Ebersberger et al., 2004; Austin et al., 2009; Ge et al., 2010; Weber et al., 2011; Hayden et al., 2012; Anderson et al., 2013). However, there seem to be consistent responses of certain bacterial and fungal taxa across experiments. For example, and in contrast with the early findings of Lipson et al., Acidobacteria have been shown to decrease in abundance under elevated CO2 (Lesaulnier et al., 2008; Dunbar et al., 2012; Hayden et al., 2012), while Actinobacteria and Bacteroidetes tend to increase (Lesaulnier et al., 2008; Nguyen

114


et al., 2011; Hayden et al., 2012). Far fewer studies report on changes in relative abundance of fungal taxa, which makes finding general patterns challenging. However, a few studies found that elevated CO2 increased the relative abundance of Basidiomycota (Lesaulnier et al., 2008; Tu et al., 2015), while Ascomycota have been shown to both increase (Nguyen et al., 2011) and decrease (Tu et al., 2015) in response to elevated CO2.

Warming While these early studies on the effects of climate change on soil microbial communities and their functioning focused on the effects of elevated CO2, later work focused primarily on changes in temperature and precipitation patterns. While elevated concentrations of atmospheric CO2 affect soil microbial communities indirectly, through altering plant physiological processes, growth, and community composition, changes in temperature can affect soil microbial communities both indirectly, through plant growth and community responses, and directly, without plant intervention. In one of the first studies on the effects of increased temperature on soil microbial communities and respiration rates, Zogg et al. (1996) subjected unvegetated soil cores to a range of temperatures and found that increased temperature resulted in an increase in some grampositive PLFAs, a decrease in gram-negative PLFAs, and no change in fungal PLFAs. Importantly, they also found that the increased respiration rates at higher temperatures were a result of increased soil C substrate pools rather than increased rate constants. Not long after that study, Kandeler et al. (1998) found, in vegetated soils, that warming increased total microbial biomass, as well as bacterial and fungal biomass, measured as PLFAs. In the same experiment, Bardgett et al. (1999) also found that gram-positive bacteria increased under warmed conditions, potentially reflecting the increase in easily degradable, labile C in the soil. However, these authors did not find increased soil respiration rates under higher temperatures. In the late 2000s, two studies came to contrasting conclusions on whether microbial respiration acclimates over time under warming or not (Bradford et al., 2008; Hartley et al., 2008). This discussion was fuelled because several field experiments showed that increased respiration rates under warming are reduced to control levels over time. Apparent thermal acclimation of soil respiration rates would have important implications for climate models, which currently predict an additional 2 C increase in temperature by 2100 because of the feedback between increased soil respiration and atmospheric CO2 concentrations. While the jury is still out on what is the dominant mechanism, this observed apparent acclimation of soil respiration rates to warming is most likely caused by a combination of substrate depletion, microbial physiological adaptation to warmer temperatures, and changes in microbial community composition (Bradford, 2013). And indeed, recent experimental evidence showed that shifts in microbial community composition, and in particular an increased abundance of gram-positive bacteria and a decrease in gram-negative bacteria and fungi,

A Short History of Research on Climate Change Impacts

115

would explain thermal acclimation of soil respiration (Wei et al., 2014). Also in the 2000s, novel DNA-based sequencing methods allowed for the more detailed quantification of the effect of warming on soil bacterial and fungal community composition, as well as on the abundances of specific microbial groups. For example, in a field experiment, Sheik et al. (2011) found that while warming increased total microbial abundance, it decreased bacterial diversity and reduced the relative abundance of Actinobacteria but increased Acidobacteria. Similarly, DeAngelis et al. (2015) found that long-term warming in forest stands increased the abundance of bacterial taxa associated with oligotrophic strategies such as Acidobacteria and Alphaproteobacteria. In other field studies, Actinobacteria as well as Alphaproteobacteria increased under warming (Deslippe et al., 2012; Hayden et al., 2012). Similar to the early studies on the effects of elevated CO2, many studies found an increase in fungal abundance under warmed conditions (Castro et al., 2010; Deslippe et al., 2012; Yergeau et al., 2012; Haugwitz et al., 2014), and these patterns have recently been confirmed in a meta-analysis (Garcia-Palacios et al., 2015). However, it is not clear whether these patterns are the result of warming directly affecting soil microbial communities or indirectly through affecting plant growth. However, recent evidence shows that plant communities strongly modify how ecosystem respiration responds to climate warming (Ward et al., 2013).

Drought It has long been recognized that soil moisture is an important driver of the composition and activity of soil communities, and the first studies to assess the effect of fluctuations in soil moisture on soil communities did not do so from a global climate change perspective. While some early studies assessed the effect of changes in soil moisture on soil faunal communities (Briones et al., 1997; Huhta et al., 1998), the first experiment to test how drought affects soil microbial communities was published in 2002 by Wilkinson et al. (2002). These authors subjected litterbags buried at two sites containing spruce and pine needles to different irrigation treatments and found that although irrigation effects on microbial communities were small, fungal PLFA was highest in the dry site with a Mediterranean climate. Taking a similar approach using litterbags, but this time within the context of global climate change, Taylor et al. (2004) found no difference in microbial biomass between constant and fluctuating irrigation treatments. Around the same time, using intact grassland soil monoliths, Griffiths et al. (2003) found that bacterial community composition, assessed by DGGE, did not change in response to drying and rewetting, despite strong physiological effects. In a study assessing the effect of a laboratorybased drought on microbial communities originating from either a dry or irrigated environment, fungi were increased in the soils from a dry environment but microbial community responses to the laboratory based drought were subtle (Williams, 2007), while in a second study by this author, irrigation increased fungal PLFA (Williams and Rice, 2007). Many later studies also found that

116


fungi, or fungal-dominated soils, are more resistant to drought than bacteria or bacterial-dominated soils, and again most of these studies used PLFA analysis (Gordon et al., 2008; Bapiri et al., 2010; De Vries et al., 2012a,b). This greater resistance of fungi to drought than bacteria has been attributed to their thicker cell walls and slower growth strategies (Schimel et al., 2007; De Vries and Shade, 2013), and, more recently, to their ability to redistribute water through their hyphal networks (Guhr et al., 2015). Similarly, gram-positive bacteria have been found to be more resistant to drought than gram-negative bacteria because of their thick cell walls, slow growth, and ability to form spores (De Vries and Shade, 2013). While studies on drought initially lagged behind studies assessing the impacts of elevated CO2 and warming on soil microbial communities, recent evidence that extreme climatic events affect ecosystem functioning more than gradual changes in atmospheric CO2 concentrations and temperature (Reichstein et al., 2013) has led to an increase in studies on drought as a climate change factor. And, as with studies on warming and elevated CO2, this increased interest in the effects of drought on soil microbial communities also resulted in an increased use of DNA- and RNA-based techniques for assessing changes in microbial communities beyond shifts in fungal and bacterial abundance. For example, in a field experiment in Wales in which summer drought was simulated for 6 years, fungal diversity, as assessed by PCR-DGGE, was reduced in drought treatments (Toberman et al., 2008). In contrast, in a study assessing the dynamics in bacterial and fungal community composition during wet-up of dry soils, fungal communities were remarkably stable while bacterial communities showed strong and consistent responses to rewetting (Barnard et al., 2013). Specifically, the bacterial phylum Acidobacteria and the class Alphaproteobacteria increased in abundance during drought, while Actinobacteria decreased but rapidly regained their initial abundance, and these patterns have been confirmed by other studies (Thomson et al., 2010; Bouskill et al., 2013; Evans and Wallenstein, 2014; Amend et al., 2016).

Other Climate Change Drivers While, as summarized, extensive research has been done on the effects of elevated CO2, warming, and drought on soil microbial community composition, the effects of other climate change drivers have only received very little attention. Yet, flooding as a result of heavy rainfall has the potential to severely affect soil microbial community composition. Several studies have found a decreased biomass of decomposer as well as AM fungi (Bossio and Scow, 1998; Drenovsky et al., 2004; Bossio et al., 2006; Mentzer et al., 2006; Unger et al., 2009), while other studies found that fungi benefitted from increased rainfall in drought-prone prairie soil (Williams, 2007; Williams and Rice, 2007). More recently, studying a natural flood event, Wagner et al. (2015) found that the biomass of fungi and gram-negative bacteria decreased strongly after flooding. These authors attributed the decline in fungi to their susceptibility to anaerobic conditions and

How Can We Predict the Effect of Climate Change on Soil Microbial

117

proposed that the subsequent increase in fungal biomass was the consequence of an increased availability in plant litter due to the flood. In addition, the lower susceptibility of gram-positive bacteria to flooding than that of gram-negative bacteria corresponds with earlier findings that this group of bacteria is more stress-resistant than gram-negative bacteria (De Vries and Shade, 2013). However, the consequences of flooding for microbial community composition remain largely unexplored. Similarly, the effects of reduced snow cover as a result of climate change on soil microbial communities have received little attention. Reduced snowfall and early snowmelt in mountain ecosystems result in reduced soil insulation and increased frequency and amplitude of freezeethaw cycles, which can strongly affect soil microbial communities. The few studies that are available point to a decrease in fungal abundance with higher soil temperatures under reduced snow cover (Robroek et al., 2013), while Aanderud et al. (2013) found that fungal abundance was reduced in the reduced snow cover treatment with strongly fluctuating temperatures. In addition, Haei et al. (2011) found that fungal growth rates were higher at subzero temperatures than bacterial growth rates. These studies corroborate earlier studies on seasonal dynamics of microbial communities in tundra and alpine systems that found a higher abundance of fungi during winter than during summer. Focusing on changes in bacterial community composition, Weedon et al. (2012) found that snow addition and spring warming reduced the abundance of Alphaproteobacteria and Betaproteobacteria in subarctic peatlands. In contrast, Kumar et al. (2013) found no consistent responses of soil bacterial communities from different Arctic locations to freezee thaw cycles. Together, these studies suggest that fungi are more resistant than bacteria to freezeethawing events, but the effects on fungal and bacterial community composition are unclear.

HOW CAN WE PREDICT THE EFFECT OF CLIMATE CHANGE ON SOIL MICROBIAL COMMUNITIES? It is clear that climate change affects soil microbial community composition, and that the different drivers affect microbial communities and the relative abundance of major microbial groups differently. From the above, a picture emerges in which the relative abundances of fungi, gram-negative bacteria, and gram-positive bacteria seem to have the most consistent responses to elevated CO2, warming, drought, flooding, and changes in snow cover. Fungi and gram-positive bacteria broadly represent K-strategists or oligotrophic organisms with slow growth rates, thick cell walls, and the ability to form spores. These characteristics have been used to explain the relatively higher stress resistance of these broad microbial groups, which, in the case of climate change, translates in a higher resistance to drought and fluctuating moisture conditions (Schimel et al., 2007; De Vries and Shade, 2013). But also, it has been suggested that the general trade-off between investing in growth and investing in

118


defence explains why K-strategists or oligotrophic microbes tend to be more resistant to disturbances (De Vries and Shade, 2013). While this idea, that oligotrophic microbial life-history strategies can explain their response to disturbances has been suggested by numerous authors, empirical evidence is accumulating to support this notion. In 2007, Fierer et al. proposed a classification of major bacterial lineages into copiotrophic and oligotrophic strategies, with Acidobacteria being oligotrophic, and Betaproteobacteria and Bacteroidetes copiotrophic (Fierer et al., 2007). Indeed, and as summarized in the previous sections, Acidobacteria have repeatedly been shown to increase in abundance in response to drought and warming, while other classes, such as Actinobacteria and taxa within the Proteobacteria, tend to decrease. And, more explicitly testing how oligotrophy versus copiotrophy affects microbial response to long-term drought, DeAngelis et al. (2015) found that drought decreased average bacterial rRNA operon copy number, which tends to correlate with copiotrophic strategies such as high maximum growth rate and the ability to change growth rates quickly. In addition, these major groups of bacteria also show consistent responses to elevated CO2, with oligotrophic Acidobacteria generally decreasing, and copiotrophic Bacteroidetes and Actinobacteria decreasing, in response to elevated CO2, confirming their respective growth strategies (Fig. 5.1). But while oligotrophic life strategies might be favored over copiotrophic ones during long-term drought and warming, and copiotrophic strategies are favored under elevated CO2, it has also been shown that after ending a drought, copiotrophic taxa return quickest to their original abundances, confirming that there is a trade-off between resistance to disturbance and maximum growth. However, while maximum growth rate can predict microbial biomass and community recovery after ending a disturbance, there are likely a multitude of other factors that can explain how a microbial community recovers, or not, after a disturbance. For example, a microbe can only realize its maximum growth rate when resource availability and moisture are not limiting. Thus, resource availability should positively affect microbial community recovery, and there is some experimental evidence for this (De Vries et al., 2012b; De Vries and Shade, 2013). But the high spatial heterogeneity in soil might also affect how, when, and where microbial communities are affected by disturbances, and whether and how fast they recover from these (Fig. 5.1). Spatial heterogeneity in soil plays an important role in shaping microbial communities, both through physically isolating metacommunities and through heterogeneity of resource availability (Zhou et al., 2002). It is not clear what the role of belowground dispersal is in microbial community recovery after disturbance, but it seems plausible to assume that factors connecting metacommunities would increase soil microbial community recovery. For example, Treves et al. (2003) showed that high moisture content facilitates dispersal between metacommunities by connecting soil pores, and it has been suggested that the movement of soil micro- and mesofauna can aid in community recovery through dispersal of microbes or their spores (De Vries and Shade, 2013).


119

Although general growth strategies might predict microbial community response to drought, warming, and elevated CO2, it seems likely that specific microbial traits enable microbial species or taxa to cope with specific disturbances (Table 5.1). For example, the production of intracellular trehalose seems to play a role in desiccation tolerance in Escherichia coli (Zhang and Van, 2012), and the synthesis of exopolysaccharides in bacteria and Archaea enables optimal growth at low temperatures (Varin et al., 2012). In contrast, press disturbances such as elevated CO2 likely select for traits that optimize the use of the resulting increased inputs of labile C (as reviewed in De Vries and Shade). While as yet, evidence that specific climate change drivers select for specific microbial traits is limited, recent evidence showing phylogenetic clustering of soil bacterial and fungal community response to climate change supports the idea that specific traits might underlie this response (Fig. 5.1). For example, the consistent response of the relative abundance of fungi compared with bacteria, and of certain bacterial taxa to drought, warming, and snow addition, as described earlier, suggests that the traits underlying the response to these climate change drivers are phylogenetically clustered. Indeed, in a recent study, the authors explicitly tested the phylogenetic clustering of the response of fungal and bacterial community composition to both nitrogen addition and drought and found that fungal community response to drought had greater phylogenetic clustering than the response to nitrogen addition, while the phylogenetic clustering of bacterial community response was equal for these disturbances (Amend et al., 2016). These authors also found that, for bacteria, multiple genes coding for the degradation of multiple C compounds were linked to the response of OTUs to nitrogen addition, while only the xylan degradation gene was linked to operational taxonomic unit (OTU) response to drought. These findings suggest that while specific microbial “response” traits might predict microbial community response to climate change, these traits do not necessarily predict the consequences for microbial functioning (Fig. 5.1). Thus, we may distinguish between microbial response traits and microbial effects traits, similar to the response-and-effect traits framework for plant community response to environmental change (Suding et al., 2008). This distinction between microbial response and effect traits is supported by a lack of studies finding direct links between changes in (taxonomic) microbial community composition and functioning as a result of climate change treatments, while a considerable number of studies do find links between changes in gene abundances and soil processes. For example, and as mentioned earlier, it is well known that warming generally increases rates of heterotrophic soil respiration (Wang et al., 2014; Garcia-Palacios et al., 2015). In addition, warming also increases the rate at which processes of N cycling are carried out, in particular nitrification and mineralization, but also potential rates of denitrification (Bai et al., 2013). While there is some evidence that these changes in process rates are linked to changes in broad taxonomic groups (Garcia-Palacios), several studies have found that underlying changes in functional genes can explain these changes. For example, warming has been shown to increase the


120

Response

Resilient

Community response to climate change

En

vir

on

me

nt

Resource-poor, isolated

Resistant Dominated by rstrategists

Microbial community structure

Dominated by Kstrategists

Resource-rich, connected

N cycling genes

Drought resistance genes

Relative abundance of specific functional genes (multidimensional space)

C cycling genes

Pulse disturbance Start

Press disturbance

End

A

Start

Process rate

Process rate

C

Time Start

Time

End

B

Start

Consequences for ecosystem functioning

Process rate

Process rate

D

Time

Time

FIGURE 5.1 Framework for predicting soil microbial community response to climate change and the consequences for soil functioning. Top: Microbial community response to climate change as determined by general growth strategies and the soil environment. Communities that are dominated by K-strategist taxa tend to be more resistant to disturbance,


121

abundance of microbial C cycling genes, and in particular those responsible for decomposing labile C compounds, and this increase in gene abundance has been linked to the observed greater respiration rates and decomposition of these C sources (Zhou et al., 2012). The same study also showed an increase in genes involved in the N cycle, and this was link to greater process rates, particularly of denitrification. Similarly, other field-based warming experiments found increases in functional genes involved in the N cycle (Cantarel et al., 2012; Yergeau et al., 2012; Billings and Tiemann, 2014; Luo et al., 2014) and C cycle (Luo et al., 2014; Xue et al., 2016), with links to process rates and soil availability of C and N. Drought is well known to slow down rates of C and N transformations, but result in a flush of C and N mineralization after rewetting (the Birch-effect, as reviewed in (Borken and Matzner, 2009). While it has been shown that these significant C and N losses can be mediated by the composition of soil food webs and soil microbial communities, and specifically by more fungaldominated soil food web and microbial communities (Gordon et al., 2008; De Vries et al., 2012a), recent evidence also shows that changes in gene abundances can explain changes in processes of C and N cycling. For example, Placella and Firestone (2013) found a rapid increase in transcripts of ammonia monooxygenases and nitrite oxidoreductase after rewetting of dry soils, and these increases were linked to a reduction in the size of the soil ammonium pool. In another study, a rainfall event simulated in two drought-stressed agricultural fields triggered a peak in soil soluble N concentrations and N2O emissions, and this peak was mirrored by a significant increase in the abundance of nitrifier and denitrifier genes (Snider et al., 2015). Similar to drought and warming, the response of microbial genes involved in the C and N cycle to elevated CO2 has been linked to the consequences for functioning. For example, in a mesocosm study in which soil with trembling

=

but recover slower than r-strategist taxa, while the soil environment can affect microbial community response through resource availability and connectedness between metacommunities. Middle: The relative abundance of specific microbial genes present in a microbial community (in multidimensional space) ultimately underlie both microbial community response to climate change, and the consequences for soil functioning. Bottom: Soil functioning responses to climate change, which might be predicted by the presence of functional genes involved in the C and N cycle. Soil processes of C and N cycling can be affected in several ways. Red arrows (dark gray in print version) indicate the initial response to a change in environment (resistance); green arrows (light gray in print version) indicate the continued response to a change of environment (recovery or resilience). A process can be reduced, and increase sharply when the disturbance is ended (for example, N mineralization under drought, A), be reduced, and not return to its initial level (for example, when a permanent change has taken place in the microbial community, B), be increased but acclimate to the new situation over time (for example, respiration under warming, C), and be increased but not acclimate over time (for example, N immobilization under elevated CO2). Figure modified from De Vries, F.T., Shade, A., 2013. Controls on soil microbial community stability under climate change. Frontiers in Microbiology 4, 265.

122


Table 5.1 Types of Extreme Events and Associated Hypothesized Microbial Response Traits Type of Extreme Event

Broad Trait Specific Category Response Traits

Possible Genes Involved

References

Drought and flooding

Desiccation tolerance, osmotolerance

Trehalose and peptidoglycan production, exopolysaccharide production, capsule O-acetylation

otsBA, otsA, murB, neuO

Mordhorst et al. (2009); Culligan et al. (2012); Zhang and Van (2012)

Heat wave and freezing

Temperature change resistance

Trehalose production, cold shock tolerance proteins: fatty acid desaturase, exopolysaccharide (EPS), and other extracellular polysaccharide synthesis, chaperone protein (DNA folding), cryoprotectant synthesis

otsBA, otsA, infA, rbfA, exopolysaccharide synthesis, fatty acid desaturases

Canovas et al. (2001); Varin et al. (2012)

Wide range of environmental stresses

Dormancy

Ability to form spores, endospores, akinetes, resuscitation factors

spo0A, spo0B, hipA-hipB

Lennon and Jones (2011); Higgins and Dworkin (2012)

Modified from De Vries, F.T., Shade, A., 2013. Controls on soil microbial community stability under climate change. Frontiers in Microbiology 4, 265.

aspen was subjected to elevated CO2, nitrite reductase (nirK) genes doubled in abundance, while bacterial ammonium monooxygenase (amoA) genes were reduced (Kelly et al., 2013). These changes in N-cycling gene abundances were linked to increased loss of N from soils subjected to elevated CO2. Similarly, Xiong et al. (2015) found that in a maize agroecosystem, elevated CO2 resulted in an increase in the abundance of C, N, and P cycling genes. In particular, genes involved in the degradation of labile and recalcitrant C were increased in abundance, and this was paralleled by a stimulation of microbial C use capacity as measured using EcoPlate. However, comparing three different soil types, Butterly et al. (2016) found that responses of genes involved in the C and N cycle to elevated CO2 were highly context dependent, and that the abundances of many functional genes were reduced in the top soil layers. But the abundances of these functional genes were correlated to soil C and N availability, and total soil C and N pools. Thus, while microbial genes involved in the C and N cycle might inform about the consequences for functioning, they might also explain the response of microbial community composition to the indirect effects of climate change drivers. This overlap in response and effects genes is most obvious for microbial

References 123 response to elevated CO2, where changes in soil availability of C and N as a result of altered plant growth and physiological processes have cascading effects on microbial community composition and functioning (de Graaff et al., 2006; Phillips et al., 2011, 2012). Therefore, we might expect that, similarly, increased plant growth under warmed conditions reduces soil N availability and thus affect soil microbial community composition and N cycling genes (Kardol et al., 2010). And after a drought has ended, the flush in availability of soil C and N might affect the abundance and diversity of microbial C and N cycling genes (Fuchslueger et al., 2014).

CONCLUSION A vast and ever growing body of research into the consequences of climate change for belowground microbial communities and the implications for ecosystem functioning draws a complicated picture. Lots of results that have been reported since the 1990s seem to be context dependent, and we sorely need to understand the role of different soil and vegetation types in driving microbial community response to climate change. However, while it might be hard to see consistent patterns, some generalities are starting to shine through. In particular, microbial groups and bacteria taxa that are associated with oligotrophic or K-strategist life-history strategies seem to be consistently increasing in abundance under drought and warming, while they decrease with elevated CO2. In contrast, under pulse disturbances such as drought followed by rewetting, the more copiotrophic or r-strategist groups, with high maximum growth rates, are able to rapidly regain their abundance (De Vries and Shade, 2013). However, specific functional genes might be more informative than broad growth strategies for predicting the response of microbial community composition to specific climate change drivers, and in particular extreme climatic events such as heat waves, droughts, flooding, and soil freezing through reduced snow cover. In addition, accumulating evidence suggests that functional genes involved in the C and N cycle can predict the consequences of changes in microbial community composition in response to climate change for soil functioning. But, specific functional genes might not only be more informative for predicting the consequences for ecosystem functioning but also for predicting the indirect effects of climate change on microbial community composition, through changes in plant growth and physiological processes and the resulting consequences for soil C and N availability.

REFERENCES Aanderud, Z.T., Jones, S.E., Schoolmaster, D.R., Fierer, N., Lennon, J.T., 2013. Sensitivity of soil respiration and microbial communities to altered snowfall. Soil Biology & Biochemistry 57, 217e227. Amend, A.S., Martiny, A.C., Allison, S.D., Berlemont, R., Goulden, M.L., Lu, Y., Treseder, K.K., Weihe, C., Martiny, J.B.H., 2016. Microbial response to simulated global change is phylogenetically conserved and linked with functional potential. ISME Journal 10, 109e118.

124


Anderson, I.C., Drigo, B., Keniry, K., Ghannoum, O., Chambers, S.M., Tissue, D.T., Cairney, J.W.G., 2013. Interactive effects of preindustrial, current and future atmospheric CO2 concentrations and temperature on soil fungi associated with two Eucalyptus species. FEMS Microbiology Ecology 83, 425e437. Anderson, J.P., Domsch, K.H., 1973. Quantification of bacterial and fungal contributions to soil respiration. Archiv Fur Mikrobiologie 93, 113e127. Austin, E.E., Castro, H.F., Sides, K.E., Schadt, C.W., Classen, A.T., 2009. Assessment of 10 years of CO2 fumigation on soil microbial communities and function in a sweetgum plantation. Soil Biology & Biochemistry 41, 514e520. Bai, E., Li, S.L., Xu, W.H., Li, W., Dai, W.W., Jiang, P., 2013. A meta-analysis of experimental warming effects on terrestrial nitrogen pools and dynamics. New Phytologist 199, 441e451. Bapiri, A., Ba˚a˚th, E., Rousk, J., 2010. Drying-rewetting cycles affect fungal and bacterial growth differently in an arable soil. Microbial Ecology 60, 419e428. Bardgett, R.D., Kandeler, E., Tscherko, D., Hobbs, P.J., Bezemer, T.M., Jones, T.H., Thompson, L.J., 1999. Below-ground microbial community development in a high temperature world. Oikos 85 (2), 193e203. Bardgett, R.D., Manning, P., Morrie¨n, E., de Vries, F.T., 2013. Hierarchical responses of plantesoil interactions to climate change: consequences for the global carbon cycle. Journal of Ecology 101, 334e343. Barnard, R.L., Osborne, C.A., Firestone, M.K., 2013. Responses of soil bacterial and fungal communities to extreme desiccation and rewetting. ISME Journal 7, 2229e2241. Billings, S.A., Tiemann, L.K., 2014. Warming-induced enhancement of soil N2O efflux linked to distinct response times of genes driving N2O production and consumption. Biogeochemistry 119, 371e386. Blankinship, J., Niklaus, P., Hungate, B., 2011. A meta-analysis of responses of soil biota to global change. Oecologia 165, 553e565. Borken, W., Matzner, E., 2009. Reappraisal of drying and wetting effects on C and N mineralization and fluxes in soils. Global Change Biology 15, 808e824. Bossio, D.A., Fleck, J.A., Scow, K.M., Fujii, R., 2006. Alteration of soil microbial communities and water quality in restored wetlands. Soil Biology and Biochemistry 38, 1223e1233. Bossio, D.A., Scow, K.M., 1998. Impacts of carbon and flooding on soil microbial communities: phospholipid fatty acid profiles and substrate utilization patterns. Microbial Ecology 35, 265e278. Bouskill, N.J., Lim, H.C., Borglin, S., Salve, R., Wood, T.E., Silver, W.L., Brodie, E.L., 2013. Pre-exposure to drought increases the resistance of tropical forest soil bacterial communities to extended drought. The ISME Journal 7, 384e394. Bradford, M.A., 2013. Thermal adaptation of decomposer communities in warming soils. Frontiers in Microbiology 4. Bradford, M.A., Davies, C.A., Frey, S.D., Maddox, T.R., Melillo, J.M., Mohan, J.E., Reynolds, J.F., Treseder, K.K., Wallenstein, M.D., 2008. Thermal adaptation of soil microbial respiration to elevated temperature. Ecology Letters 11, 1316e1327. Briones, M.J.I., Ineson, P., Piearce, T.G., 1997. Effects of climate change on soil fauna; Responses of enchytraeids, Diptera larvae and tardigrades in a transplant experiment. Applied Soil Ecology 6, 117e134. Butterly, C.R., Phillips, L.A., Wiltshire, J.L., Franks, A.E., Armstrong, R.D., Chen, D.L., Mele, P.M., Tang, C.X., 2016. Long-term effects of elevated CO2 on carbon and nitrogen functional capacity of microbial communities in three contrasting soils. Soil Biology & Biochemistry 97, 157e167. Canovas, D., Fletcher, S.A., Hayashi, M., Csonka, L.N., 2001. Role of trehalose in growth at high temperature of Salmonella enterica serovar typhimurium. Journal of Bacteriology 183, 3365e3371. Cantarel, A.A.M., Bloor, J.M.G., Pommier, T., Guillaumaud, N., Moirot, C., Soussana, J.-F., Poly, F., 2012. Four years of experimental climate change modifies the microbial drivers of N2O fluxes in an upland grassland ecosystem. Global Change Biology 18, 2520e2531. Castro, H.F., Classen, A.T., Austin, E.E., Norby, R.J., Schadt, C.W., 2010. Soil microbial community responses to multiple experimental climate change drivers. Applied and Environmental Microbiology 76, 999e1007.

References 125 Cheng, L., Booker, F.L., Burkey, K.O., Tu, C., Shew, H.D., Rufty, T.W., Fiscus, E.L., Deforest, J.L., Hu, S.J., 2011. Soil microbial responses to elevated CO2 and O-3 in a nitrogen-aggrading agroecosystem. PLoS One 6. Culligan, E.P., Sleator, R.D., Marchesi, J.R., Hill, C., 2012. Functional metagenomics reveals novel salt tolerance loci from the human gut microbiome. ISME Journal 6, 1916e1925. de Graaff, M.A., van Groenigen, K.J., Six, J., Hungate, B., van Kessel, C., 2006. Interactions between plant growth and soil nutrient cycling under elevated CO2: a meta-analysis. Global Change Biology 12, 2077e2091. De Vries, F.T., Liiri, M., Bjørnlund, L., Bowker, M., Christensen, S., Seta¨la¨, H., Bardgett, R.D., 2012a. Land use alters the resistance and resilience of soil food webs to drought. Nature Climate Change 2, 276e280. De Vries, F.T., Liiri, M., Bjørnlund, L., Seta¨la¨, H., Christensen, S., Bardgett, R.D., 2012b. Legacy effects of drought on plant growth and the soil food web. Oecologia 170, 821e833. De Vries, F.T., Shade, A., 2013. Controls on soil microbial community stability under climate change. Frontiers in Microbiology 4, 265. DeAngelis, K.M., Pold, G., Topcuoglu, B.D., van Diepen, L.T.A., Varney, R.M., Blanchard, J.L., Melillo, J., Frey, S.D., 2015. Long-term forest soil warming alters microbial communities in temperate forest soils. Frontiers in Microbiology 6. Deslippe, J.R., Hartmann, M., Simard, S.W., Mohn, W.W., 2012. Long-term warming alters the composition of Arctic soil microbial communities. FEMS Microbiology Ecology 82, 303e315. Drenovsky, R.E., Vo, D., Graham, K.J., Scow, K.M., 2004. Soil water content and organic carbon availability are major determinants of soil microbial community composition. Microbial Ecology 48, 424e430. Drigo, B., Kowalchuk, G.A., Yergeau, E., Bezemer, T.M., Boschker, H.T.S., Van Veen, J.A., 2007. Impact of elevated carbon dioxide on the rhizosphere communities of Carex arenaria and Festuca rubra. Global Change Biology 13 (11), 2396e2410. Dunbar, J., Eichorst, S.A., Gallegos-Graves, L., Silva, S., Xie, G., Hengartner, N.W., Evans, R.D., Hungate, B.A., Jackson, R.B., Megonigal, J.P., Schadt, C.W., Vilgalys, R., Zak, D.R., Kuske, C.R., 2012. Common bacterial responses in six ecosystems exposed to 10 years of elevated atmospheric carbon dioxide. Environmental Microbiology 14, 1145e1158. Ebersberger, D., Werrnbter, N., Niklaus, P.A., Kandeler, E., 2004. Effects of long term CO2 enrichment on microbial community structure in calcareous grassland. Plant and Soil 264, 313e323. Evans, S.E., Wallenstein, M.D., 2014. Climate change alters ecological strategies of soil bacteria. Ecology Letters 17, 155e164. Fierer, N., Bradford, M.A., Jackson, R.B., 2007. Toward an ecological classification of soil bacteria. Ecology 88, 1354e1364. ˚ ., Tunlid, A., Ba˚a˚th, E., 1993. Phospholipid fatty acid composition, biomass, and Frostega˚rd, A activity of microbial communities from two soil types experimentally exposed to different heavy metals. Applied and Environmental Microbiology 59, 3605e3617. Fuchslueger, L., Kastl, E.M., Bauer, F., Kienzl, S., Hasibeder, R., Ladreiter-Knauss, T., Schmitt, M., Bahn, M., Schloter, M., Richter, A., Szukics, U., 2014. Effects of drought on nitrogen turnover and abundances of ammonia-oxidizers in mountain grassland. Biogeosciences 11, 6003e6015. Garcia-Palacios, P., Vandegehuchte, M.L., Shaw, E.A., Dam, M., Post, K.H., Ramirez, K.S., Sylvain, Z.A., de Tomasel, C.M., Wall, D.H., 2015. Are there links between responses of soil microbes and ecosystem functioning to elevated CO2, N deposition and warming? A global perspective. Global Change Biology 21, 1590e1600. Ge, Y.A., Chen, C.R., Xu, Z.H., Oren, R., He, J.Z., 2010. The spatial factor, rather than elevated CO2, controls the soil bacterial community in a temperate forest ecosystem. Applied and Environmental Microbiology 76, 7429e7436. Gordon, H., Haygarth, P.M., Bardgett, R.D., 2008. Drying and rewetting effects on soil microbial community composition and nutrient leaching. Soil Biology & Biochemistry 40, 302e311.

126


Griffiths, R.I., Whiteley, A.S., O’Donnell, A.G., Bailey, M.J., 2003. Physiological and community responses of established grassland bacterial populations to water stress. Applied and Environmental Microbiology 69, 6961e6968. Guenet, B., Lenhart, K., Leloup, J., Giusti-Miller, S., Pouteau, V., Mora, P., Nunan, N., Abbadie, L., 2012. The impact of long-term CO2 enrichment and moisture levels on soil microbial community structure and enzyme activities. Geoderma 170, 331e336. Guhr, A., Borken, W., Spohn, M., Matzner, E., 2015. Redistribution of soil water by a saprotrophic fungus enhances carbon mineralization. Proceedings of the National Academy of Sciences 112, 14647e14651. Haei, M., Rousk, J., Ilstedt, U., Oquist, M., Baath, E., Laudon, H., 2011. Effects of soil frost on growth, composition and respiration of the soil microbial decomposer community. Soil Biology & Biochemistry 43, 2069e2077. Hartley, I.P., Hopkins, D.W., Garnett, M.H., Sommerkorn, M., Wookey, P.A., 2008. Soil microbial respiration in arctic soil does not acclimate to temperature. Ecology Letters 11, 1092e1100. Haugwitz, M.S., Bergmark, L., Prieme, A., Christensen, S., Beier, C., Michelsen, A., 2014. Soil microorganisms respond to five years of climate change manipulations and elevated atmospheric CO2 in a temperate heath ecosystem. Plant and Soil 374, 211e222. Hayden, H.L., Mele, P.M., Bougoure, D.S., Allan, C.Y., Norng, S., Piceno, Y.M., Brodie, E.L., DeSantis, T.Z., Andersen, G.L., Williams, A.L., Hovenden, M.J., 2012. Changes in the microbial community structure of bacteria, archaea and fungi in response to elevated CO2 and warming in an Australian native grassland soil. Environmental Microbiology 14, 3081e3096. He, Z.L., Xu, M.Y., Deng, Y., Kang, S.H., Kellogg, L., Wu, L.Y., Van Nostrand, J.D., Hobbie, S.E., Reich, P.B., Zhou, J.Z., 2010. Metagenomic analysis reveals a marked divergence in the structure of belowground microbial communities at elevated CO2. Ecology Letters 13, 564e575. Higgins, D., Dworkin, J., 2012. Recent progress in Bacillus subtilis sporulation. FEMS Microbiology Reviews 36, 131e148. Huhta, V., Sulkava, P., Viberg, K., 1998. Interactions between enchytraeid (Cognettia sphagnetorum), microarthropod and nematode populations in forest soil at different moistures. Applied Soil Ecology 9, 53e58. IPCC, 2014. Summary for policymakers. In: Christopher B. Field (USA), V.R.B.A., Michael D. Mastrandrea (USA), Katharine J. Mach (USA), M.A.-K.A.E., W. Neil Adger (UK), Yury A., Anokhin (Russian Federation), O.A.A.R.F., Douglas J. Arent (USA), Jonathon Barnett (Australia), V.R.B.U., Rongshuo Cai (China), Monalisa, Chatterjee (USA/India), S.J.C.C., Wolfgang Cramer (Germany/France), Purnamita Dasgupta (India), D.J.D.C., Fatima Denton (Gambia), Petra Do¨ll, (Germany), K.D.U., Yasuaki Hijioka (Japan), Ove Hoegh-Guldberg (Australia), Richard G. Jones (UK), R.N.J.A., Roger L. Kitching (Australia), R. Sari, Kovats (UK), P.R.L.M., Joan Nymand Larsen (Iceland), Erda Lin, (China), D.B.L.U., In˜igo J. Losada (Spain), Graciela O. Magrin (Argentina), Jose´, A. Marengo (Brazil), A.M.S., Bruce A. McCarl (USA), Roger F. McLean, (Australia), L.O.M.U., Guy F. Midgley (South Africa), Nobuo Mimura (Japan), John F. Morton (UK), I.N.S., Ian R. Noble (Australia), Leonard A. Nurse, (Barbados), K.L.O.B.N., Taikan Oki (Japan), Lennart Olsson (Sweden), Michael, Oppenheimer (USA), J.T.O.U., Joy J. Pereira (Malaysia), Elvira S., Poloczanska (Australia), J.R.P.D., Hans-O. Po¨rtner (Germany), Michael J., Prather (USA), R.S.P.U., Andy R. Reisinger (New Zealand), Aromar Revi, (India), O.C.R.N., David E. Satterthwaite (UK), Daniela N. Schmidt (UK), Josef Settele (Germany), K.R.S.U., Daíthı´ A. Stone (Canada/South Africa/USA), Avelino G. Suarez (Cuba), P.T.U., Riccardo Valentini (Italy), Alicia Villamizar, (Venezuela), R.W.U., Thomas J. Wilbanks (USA), Poh Poh Wong (Singapore), Alistair Woodward (New Zealand), G.W.Y.U. (Eds.), Climate Change 2014: Impacts, Adaptation, and Vulnerability. Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA.

References 127 Jones, T.H., Thompson, L.J., Lawton, J.H., Bezemer, T.M., Bardgett, R.D., Blackburn, T.M., Bruce, K.D., Cannon, P.F., Hall, G.S., Hartley, S.E., Howson, G., Jones, C.G., Kampichler, C., Kandeler, E., Ritchie, D.A., 1998. Impacts of rising atmospheric carbon dioxide on model terrestrial ecosystems. Science 280 (5362), 441e443. Kandeler, E., Tscherko, D., Bardgett, R.D., Hobbs, P.J., Kampichler, C., Jones, T.H., 1998. The response of soil microorganisms and roots to elevated CO2 and temperature in a terrestrial model ecosystem. Plant and Soil 202 (2), 251e262. Kardol, P., Campany, C.E., Souza, L., Norby, R.J., Weltzin, J.F., Classen, A.T., 2010. Climate change effects on plant biomass alter dominance patterns and community evenness in an experimental old-field ecosystem. Global Change Biology 16, 2676e2687. Kelly, J.J., Peterson, E., Winkelman, J., Walter, T.J., Rier, S.T., Tuchman, N.C., 2013. Elevated atmospheric CO2 impacts abundance and diversity of nitrogen cycling functional genes in soil. Microbial Ecology 65, 394e404. Klironomos, J.N., Rillig, M.C., Allen, M.F., 1996. Below-ground microbial and microfaunal responses to Artemisia tridentata grown under elevated atmospheric CO2. Functional Ecology 10, 527e534. Kumar, N., Grogan, P., Chu, H., Christiansen, C.T., Walker, V.K., 2013. The effect of freezethaw conditions on arctic soil bacterial communities. Biology 2, 356e377. Lennon, J.T., Jones, S.E., 2011. Microbial seed banks: the ecological and evolutionary implications of dormancy. Nature Reviews Microbiology 9, 119e130. Lesaulnier, C., Papamichail, D., McCorkle, S., Ollivier, B., Skiena, S., Taghavi, S., Zak, D., van der Lelie, D., 2008. Elevated atmospheric CO2 affects soil microbial diversity associated with trembling aspen. Environmental Microbiology 10, 926e941. Lipson, D.A., Blair, M., Barron-Gafford, G., Grieve, K., Murthy, R., 2006. Relationships between microbial community structure and soil processes under elevated atmospheric carbon dioxide. Microbial Ecology 51, 302e314. Luo, C.W., Rodriguez, R.L.M., Johnston, E.R., Wu, L.Y., Cheng, L., Xue, K., Tu, Q.C., Deng, Y., He, Z.L., Shi, J.Z., Yuan, M.M., Sherry, R.A., Li, D.J., Luo, Y.Q., Schuur, E.A.G., Chain, P., Tiedje, J.M., Zhou, J.Z., Konstantinidis, K.T., 2014. Soil microbial community responses to a decade of warming as revealed by comparative metagenomics. Applied and Environmental Microbiology 80, 1777e1786. Mentzer, J.L., Goodman, R.M., Balser, T.C., 2006. Microbial response over time to hydrologic and fertilization treatments in a simulated wet prairie. Plant and Soil 284, 85e100. Mordhorst, I.L., Claus, H., Ewers, C., Lappann, M., Schoen, C., Elias, J., Batzilla, J., Dobrindt, U., Wieler, L.H., Bergfeld, A.K., Muhlenhoff, M., Vogel, U., 2009. O-acetyltransferase gene neuO is segregated according to phylogenetic background and contributes to environmental desiccation resistance in Escherichia coli K1. Environmental Microbiology 11, 3154e3165. Nakas, J.P., Klein, D.A., 1979. Decomposition of microbial cell components in a semi-arid grassland soil. Applied and Environmental Microbiology 38, 454e460. Nakas, J.P., Klein, D.A., 1980. Mineralization capacity of bacteria and fungi from the rhizosphere-rhizoplane of a semi-arid grassland. Applied and Environmental Microbiology 39, 113e117. Newton, P.C.D., Clark, H., Bell, C.C., Glasgow, E.M., Tate, K.R., Ross, D.J., Yeates, G.W., Saggar, S., 1995. Plant growth and soil processes in temperate grassland communities at elevated CO2. Journal of Biogeography 22, 235e240. Nguyen, L.M., Buttner, M.P., Cruz, P., Smith, S.D., Robleto, E.A., 2011. Effects of elevated atmospheric CO2 on rhizosphere soil microbial communities in a Mojave Desert ecosystem. Journal of Arid Environments 75, 917e925. Phillips, R.P., Finzi, A.C., Bernhardt, E.S., 2011. Enhanced root exudation induces microbial feedbacks to N cycling in a pine forest under long-term CO2 fumigation. Ecology Letters 14, 187e194. Phillips, R.P., Meier, I.C., Bernhardt, E.S., Grandy, A.S., Wickings, K., Finzi, A.C., 2012. Roots and fungi accelerate carbon and nitrogen cycling in forests exposed to elevated CO2. Ecology Letters 15, 1042e1049.

128


Placella, S.A., Firestone, M.K., 2013. Transcriptional response of nitrifying communities to wetting of dry soil. Applied and Environmental Microbiology 79, 3294e3302. Raich, J.W., Schlesinger, W.H., 1992. The global carbon-dioxide flux in soil respiration and its relationship to vegetation and climate. Tellus Series B-Chemical and Physical Meteorology 44, 81e99. Reichstein, M., Bahn, M., Ciais, P., Frank, D., Mahecha, M.D., Seneviratne, S.I., Zscheischler, J., Beer, C., Buchmann, N., Frank, D.C., Papale, D., Rammig, A., Smith, P., Thonicke, K., van der Velde, M., Vicca, S., Walz, A., Wattenbach, M., 2013. Climate extremes and the carbon cycle. Nature 500, 287e295. Robroek, B.J.M., Heijboer, A., Jassey, V.E.J., Hefting, M.M., Rouwenhorst, T.G., Buttler, A., Bragazza, L., 2013. Snow cover manipulation effects on microbial community structure and soil chemistry in a mountain bog. Plant and Soil 369, 151e164. Schimel, J., Balser, T.C., Wallenstein, M., 2007. Microbial stress-response physiology and its implications for ecosystem function. Ecology 88, 1386e1394. Schnurer, J., Clarholm, M., Bostrom, S., Rosswall, T., 1986. Effects of moisture on soilmicroorganisms and nematodes - a field experiment. Microbial Ecology 12, 217e230. Sheik, C.S., Beasley, W.H., Elshahed, M.S., Zhou, X.H., Luo, Y.Q., Krumholz, L.R., 2011. Effect of warming and drought on grassland microbial communities. ISME Journal 5, 1692e1700. Snider, D., Thompson, K., Wagner-Riddle, C., Spoelstra, J., Dunfield, K., 2015. Molecular techniques and stable isotope ratios at natural abundance give complementary inferences about N2O production pathways in an agricultural soil following a rainfall event. Soil Biology & Biochemistry 88, 197e213. Suding, K.N., Lavorel, S., Chapin, F.S., Cornelissen, J.H.C., Diaz, S., Garnier, E., Goldberg, D., Hooper, D.U., Jackson, S.T., Navas, M.L., 2008. Scaling environmental change through the community-level: a trait-based response-and-effect framework for plants. Global Change Biology 14, 1125e1140. Taylor, A.R., Schroter, D., Pflug, A., Wolters, V., 2004. Response of different decomposer communities to the manipulation of moisture availability: potential effects of changing precipitation patterns. Global Change Biology 10, 1313e1324. Thomson, B.C., Ostle, N.J., McNamara, N.P., Whiteley, A.S., Griffiths, R.I., 2010. Effects of sieving, drying and rewetting upon soil bacterial community structure and respiration rates. Journal of Microbiological Methods 83, 69e73. Toberman, H., Freeman, C., Evans, C., Fenner, N., Artz, R.R.E., 2008. Summer drought decreases soil fungal diversity and associated phenol oxidase activity in upland Calluna heathland soil. FEMS Microbiology Ecology 66, 426e436. Treves, D.S., Xia, B., Zhou, J., Tiedje, J.M., 2003. A two-species test of the hypothesis that spatial isolation influences microbial diversity in soil. Microbial Ecology 45, 20e28. Tu, Q., Yuan, M., He, Z., Deng, Y., Xue, K., Wu, L., Hobbie, S.E., Reich, P.B., Zhou, J., 2015. Fungal communities respond to long-term CO2 elevation by community reassembly. Applied and Environmental Microbiology 81, 2445e2454. Unger, I.M., Kennedy, A.C., Muzika, R.M., 2009. Flooding effects on soil microbial communities. Applied Soil Ecology 42, 1e8. Van der Heijden, M.G.A., Bardgett, R.D., van Straalen, N.M., 2008. The unseen majority: soil microbes as drivers of plant diversity and productivity in terrestrial ecosystems. Ecology Letters 11, 296e310. Van Groenigen, K.J., Six, J., Harris, D., Van Kessel, C., 2007. Elevated CO2 does not favor a fungal decomposition pathway. Soil Biology & Biochemistry 39, 2168e2172. Varin, T., Lovejoy, C., Jungblut, A.D., Vincent, W.F., Corbeil, J., 2012. Metagenomic analysis of stress genes in microbial mat communities from Antarctica and the High Arctic. Applied and Environmental Microbiology 78, 549e559. Wagner, D., Eisenhauer, N., Cesarz, S., 2015. Plant species richness does not attenuate responses of soil microbial and nematode communities to a flood event. Soil Biology & Biochemistry 89, 135e149.

References 129 Wang, X., Liu, L.L., Piao, S.L., Janssens, I.A., Tang, J.W., Liu, W.X., Chi, Y.G., Wang, J., Xu, S., 2014. Soil respiration under climate warming: differential response of heterotrophic and autotrophic respiration. Global Change Biology 20, 3229e3237. Ward, S.E., Ostle, N.J., Oakley, S., Quirk, H., Henrys, P.A., Bardgett, R.D., 2013. Warming effects on greenhouse gas fluxes in peatlands are modulated by vegetation composition. Ecology Letters 16, 1285e1293. Weber, C.F., Zak, D.R., Hungate, B.A., Jackson, R.B., Vilgalys, R., Evans, R.D., Schadt, C.W., Megonigal, J.P., Kuske, C.R., 2011. Responses of soil cellulolytic fungal communities to elevated atmospheric CO2 are complex and variable across five ecosystems. Environmental Microbiology 13, 2778e2793. Weedon, J.T., Kowalchuk, G.A., Aerts, R., van Hal, J., van Logtestijn, R., Tas, N., Roling, W.F.M., van Bodegom, P.M., 2012. Summer warming accelerates sub-arctic peatland nitrogen cycling without changing enzyme pools or microbial community structure. Global Change Biology 18, 138e150. Wei, H., Guenet, B., Vicca, S., Nunan, N., AbdElgawad, H., Pouteau, V., Shen, W., Janssens, I.A., 2014. Thermal acclimation of organic matter decomposition in an artificial forest soil is related to shifts in microbial community structure. Soil Biology and Biochemistry 71, 1e12. Wilkinson, S.C., Anderson, J.M., Scardelis, S.P., Tisiafouli, M., Taylor, A., Wolters, V., 2002. PLFA profiles of microbial communities in decomposing conifer litters subject to moisture stress. Soil Biology & Biochemistry 34, 189e200. Williams, M.A., 2007. Response of microbial communities to water stress in irrigated and drought-prone tallgrass prairie soils. Soil Biology & Biochemistry 39, 2750e2757. Williams, M.A., Rice, C.W., 2007. Seven years of enhanced water availability influences the physiological, structural, and functional attributes of a soil microbial community. Applied Soil Ecology 35, 535e545. Xiong, J.B., He, Z.L., Shi, S.J., Kent, A., Deng, Y., Wu, L.Y., Van Nostrand, J.D., Zhou, J.Z., 2015. Elevated CO2 shifts the functional structure and metabolic potentials of soil microbial communities in a C-4 agroecosystem. Scientific Reports 5. Xue, K., Xie, J.P., Zhou, A.F., Liu, F.F., Li, D.J., Wu, L.Y., Deng, Y., He, Z.L., Van Nostrand, J.D., Luo, Y.Q., Zhou, J.Z., 2016. Warming alters expressions of microbial functional genes important to ecosystem functioning. Frontiers in Microbiology 7. Yergeau, E., Bokhorst, S., Kang, S., Zhou, J., Greer, C.W., Aerts, R., Kowalchuk, G.A., 2012. Shifts in soil microorganisms in response to warming are consistent across a range of Antarctic environments. ISME Journal 6, 692e702. Zak, D.R., Pregitzer, K.S., Curtis, P.S., Holmes, W.E., 2000. Atmospheric CO2 and the composition and function of soil microbial communities. Ecological Applications 10, 47e59. Zak, D.R., Pregitzer, K.S., Curtis, P.S., Teeri, J.A., Fogel, R., Randlett, D.L., 1993. Elevated atmospheric CO2 and feedback between carbon and nitrogen cycles. Plant and Soil 151, 105e117. Zak, D.R., Ringelberg, D.B., Pregitzer, K.S., Randlett, D.L., White, D.C., Curtis, P.S., 1996. Soil microbial communities beneath Populus grandidentata crown under elevated atmospheric CO2. Ecological Applications 6, 257e262. Zhang, Q., Van, T., 2012. Correlation of intracellular trehalose concentration with desiccation resistance of soil Escherichia coli populations. Applied and Environmental Microbiology 78, 7407e7413. Zhou, J.Z., Xia, B.C., Treves, D.S., Wu, L.Y., Marsh, T.L., O’Neill, R.V., Palumbo, A.V., Tiedje, J.M., 2002. Spatial and resource factors influencing high microbial diversity in soil. Applied and Environmental Microbiology 68, 326e334. Zhou, J.Z., Xue, K., Xie, J.P., Deng, Y., Wu, L.Y., Cheng, X.H., Fei, S.F., Deng, S.P., He, Z.L., Van Nostrand, J.D., Luo, Y.Q., 2012. Microbial mediation of carbon-cycle feedbacks to climate warming. Nature Climate Change 2 (2), 106e110. Zogg, G.P., Zak, D.R., Ringelberg, D.B., MacDonald, N.W., Pregitzer, K.S., White, D.C., 1997. Compositional and functional shifts in microbial communities due to soil warming. Soil Science Society of America Journal 61 (2), 475e481.


Chapter | Six

Nitrous Oxide Production From Soils in the Future: Processes, Controls, and Responses to Climate Change

Xia Zhu-Barker*, 1, Kerri L. Steenwerthx

*Department of Land, Air and Water Resource, University of California, Davis, CA, United States; xUSDA-ARS, Crops Pathology and Genetics Research Unit, Department of Viticulture and Enology, University of California, Davis, CA, United States 1


INTRODUCTION Nitrous oxide (N2O) is a potent greenhouse gas (GHG) contributing to positive radiative forcing and ozone destruction in the stratosphere (Dickinson and Cicerone, 1986; Stein and Yung, 2003; Ravishankara et al., 2009; Montzka et al., 2011). It has an atmospheric lifetime of approximately 114 years (IPCC, 2007c), with a 100-year global warming potential 298 times that of carbon dioxide (CO2) (IPCC, 2013). Atmospheric N2O concentrations have increased by approximately 20% since preindustrial times, with an average increase of 0.75 ppb/yr since 1970 (IPCC, 2014). Globally, soils, sediments, and water bodies are the main sources of N2O emissions, with agricultural and natural soils together representing approximately 48% of all the global N2O emission sources (Syakila and Kroeze, 2011; IPCC, 2013) (Fig. 6.1). In soils, N2O is produced from enzymatic processes mediated by microbes and fungi, including nitrification, nitrifier denitrification, and denitrification 131 Climate Change Impacts on Soil Processes and Ecosystem Properties. https://doi.org/10.1016/B978-0-444-63865-6.00006-5 Copyright © 2018 Elsevier B.V. All rights reserved.

132

Nitrous Oxide Production From Soils in the Future

Agricultural soil Natural soil Ocean Industry Deposition Animal production Atmopsheric chemistry Biomass&biofuel burning Rivers, estuaries, coastal zones

FIGURE 6.1 Relative proportions of total global nitrous oxide emissions from various sources. Adapted from data in IPCC, 2013. Anthropogenic and natural radiative forcing. In: Stocker, T.F., Qin, D., Plattner, G.-K., Tignor, M., Allen, S.K., Boschung, J., Nauels, A., Xia, Y., Bex, V., Midgley, P.M. (Eds.), Climate Change 2013: The Physical Science Basis. Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA; Syakila, A., Kroeze, C., 2011. The global nitrous oxide budget revisited. Greenhouse Gas Measurement and Management 1, 17e26.

(Robertson and Tiedje, 1987; Firestone and Davidson, 1989a; Conrad, 1996; Wrage et al., 2001; Stein, 2011; Schreiber et al., 2012; Thomson et al., 2012; Zhu et al., 2013a). Abiotic processes contribute to N2O production as well, including hydroxylamine decomposition and chemodenitrification (Chalk and Smith, 1983; Van Cleemput, 1998; Liu et al., 2014; Zhu-Barker et al., 2015a). The factors that influence these processesdpH, organic matter, oxygen (O2) availability, soil water content, soil texture, and supply of inorganic nitrogen (N)dare often assumed to be the same factors that regulate N2O production (Linn and Doran, 1984; Williams et al., 1992; Stark and Firestone, 1995; Stevens et al., 1998; Azam et al., 2002; De Bie et al., 2002; Zhu et al., 2013a; Venkiteswaran et al., 2014). Nitrous oxide is also consumed in soil through both microbial and chemical processes; thus, the amount of N2O in the atmosphere at any given time and its exchange between soil and atmosphere reflect the equilibrium between production and consumption. In general, N2O consumption requires strictly anoxic conditions, which are rarely found in cropping systems except for rice (Ye and Horwath, 2016). Although gross N2O consumption in soil systems has been recognized for a long time (Firestone and Davidson, 1989a), limited information is available to date (Chapuis-Lardy et al., 2007; Qin et al., 2014). Most studies have focused on the net production of N2O because it is usually much higher than consumption and have neglected the consumption process due to the challenge in quantifying consumption (Yang et al., 2011). Therefore, this chapter only focuses on N2O production processes.

Biological Processes that Produce N2O in Soils

133

The application of N to soils to produce more food on a limited land area has accelerated the global N cycle and, consequently, led to increases in N2O emitted to the atmosphere. There is little doubt that the use of N fertilizer and manure in soils is the largest contributor to the increase in atmospheric N2O in recent times (Bouwman et al., 2002a; Stehfest and Bouwman, 2006; IPCC, 2007b; Davidson, 2009). Other land management practices such as tillage, irrigation, and cover cropping can also influence soil N2O production and consumption processes (Lee et al., 2006; Halvorson et al., 2010b; Kallenbach et al., 2010; Venterea et al., 2011). Efforts to reduce N2O emissions from soils require sound information regarding the processes that lead to N2O production and how management practices impact or even can be used to control these processes. Climate change may affect soil N2O emissions on both short and long timescales (Gruber and Galloway, 2008; Schilt et al., 2010). The future climate is projected to have changes in temperature and hydrology regimes with increasing frequency in extreme weather events, leading to potential shifts in land use type and intensifying the need for mitigation and adaptation strategies for agriculture. While this description is clearly overly simplified, the factors outlined here can influence soil N2O emissions (Skiba and Smith, 2000; Bouwman et al., 2002b; Galloway et al., 2008; Gruber and Galloway, 2008). Thus, the response of management practices that reduce N2O emissions and its atmospheric concentrations must be evaluated for how well they perform under present conditions and future climate analogs. In many instances, mitigation of GHG emissions by agriculture necessarily will fall second to adaptation to a changing climate to achieve the imperative of improving food security in a changing climate. Still, mitigation of GHG emissions can be a cobenefit of adaptation. This chapter provides a summary of the current knowledge concerning biotic and abiotic processes that lead to the production of N2O in soils. We present how agricultural management practices control N2O emissions through their influence on these biotic and abiotic processes, and we limit the evaluation of the role of these processes leading to N2O emissions to mitigation. We also present content on how agricultural practices that can be incorporated into climate change mitigation strategies in agriculture. We intend for this review to facilitate the design of future experiments to quantify specific N2O production processes and to integrate options for mitigation, adaptation, and crop productivity through the development of predictive models.

BIOLOGICAL PROCESSES THAT PRODUCE N2O IN SOILS The biological processes that generate N2O in soils have been regarded traditionally as either nitrification or denitrification (Robertson and Tiedje, 1987; Firestone and Davidson, 1989a). Nitrification is an aerobic process in which ammonium ðNH4 þ Þ is oxidized to nitrate ðNO3 Þ. The first step of nitrification

134


FIGURE 6.2 Main biological processes of nitrous oxide (N2O) in soils and the metal content in the gene-coding enzymes that carry out these processes. Each circle represents one metal atom. Parentheses show varying metal content of a given enzyme. Solid colored squares indicate that an unknown amount of metal is present. Crosses over an arrow indicate uncertain processes. AMO, ammonia monooxygenase; cNOR, nitric oxide reductase; HAO, hydroxylamine oxidoreductase; NAR: dissimilatory nitrate reductase; NirK, Cunitrite reductase; Nos, nitrous oxide reductase; Nirs, Fe-nitrite reductase; NXR, nitrite oxidoreductase. Depictions of the N2O production pathways and metal content in each enzyme are adapted from Zhu et al. (2013) and Glass and Orphan (2012), respectively.

is the oxidation of ammonia with ammonia monooxygenase to nitrite ðNO2 Þ via hydroxylamine (NH2OH) (Fig. 6.2). This biotic process, performed by chemoautotrophic nitrifiers using ammonia as an energy source (Hollocher et al., 1981), can produce N2O by several pathways. These pathways are namely nitrifier nitrification, nitrifier denitrification, and nitrification-coupled denitrification (Hooper and Terry, 1979; Wrage et al., 2001; Zhu et al., 2013a) (Fig. 6.2). As the first step of these three pathways is ammonia oxidation, a process that can be inhibited by the well-know inhibitor acetylene (C2H2) at concentrations between 0.1 and 10 Pa, Zhu et al. (2013a) adopted the term ammonia oxidation pathways to collectively describe these three pathways. Denitrification, also referred to as heterotrophic denitrification, is performed by heterotrophic bacteria using nitrate ðNO3 Þ or NO2 as alternate electron acceptors to O2. Nitrous oxide is produced in this process as an intermediate. Other biological processes, such as heterotrophic nitrification, fungal mediated denitrification, Feammox, comammox, and dissimilatory NO3 reduction to NH4 þ (DNRA), can contribute to N2O production in soils as well (Smith, 1982; Tiedje, 1988; Papen et al., 1989; Robertson and Kuenen, 1990; Shoun et al., 1992; Anderson et al., 1993; Stevens and Laughlin, 1998; Laughlin and Stevens, 2002; Cle´ment et al., 2005; Zhang et al., 2011).

Ammonia Oxidation Pathways

135

AMMONIA OXIDATION PATHWAYS Nitrifier Nitrification Nitrifier nitrification is the oxidation of ammonia to nitrate ðNO3 Þ via nitrite (NO2 ) by nitrifiers. Two groups of microorganisms are involved in this process: the ammonia oxidizers are the primary nitrifiers that carry out the first step, while NO2 oxidizers are the secondary nitrifiers that oxidize NO2 to NO3 (Bock et al., 1986) (Fig. 6.2). The biotic decomposition of NH2OH, the first intermediate of nitrification, is carried out by hydroxylamine oxidoreductase, an enzyme which catalyzes the oxidation of hydroxylamine to form NO; limited evidence suggests that N2O can also be released during this process (Hooper and Terry, 1979). As hydroxylamine can also be decomposed through chemical reactions with redox-active metals (e.g., iron(III) and manganese(IV)), it is often difficult to determine whether N2O is generated via nitrifier nitrification or abiotic reactions as further described in the following section.

Nitrifier Denitrification Nitrifier denitrification is the process by which autotrophic ammonia oxidizers reduce NO2 , produced from the oxidation of ammonia, to N2O and potentially N2 (Poth, 1986; Muller et al., 1995); this occurs under limited O2 conditions (Fig. 6.2). The first part of this pathway is attributed to nitrification, whereas the reduction of NO2 is regarded as denitrification (Poth and Focht, 1985; Muller et al., 1995). The organisms involved in nitrifier denitrification are mostly ammonia-oxidizing bacteria (AOB) (Kuai and Verstraete, 1998). Ammonia-oxidizing archaea (AOA), a group of ammonia oxidizers that has been found to numerically outcompete AOB in soils (Leininger et al., 2006), are unlikely to participate in nitrifier denitrification because they lack genes encoding a potential NO reductase to be involved in nitrifier denitrification (Walker et al., 2010; Tourna et al., 2011). The NO2 reductase required by nitrifier denitrification was first characterized by Hooper (1968) in the cell extracts of Nitrosomonas europaea after incubating cells in a solution containing 15NNO2 plus ammonia under an atmosphere of 0.1% O2. The capacity for denitrification in autotrophic nitrifiers has been consistently documented in pure culture since its discovery (Ritchie and Nicholas, 1972; Goreau et al., 1980; Lipschultz et al., 1981; Poth and Focht, 1985; Anderson and Levine, 1986; Anderson et al., 1993; Shaw et al., 2006). It has been suggested that nitrifier denitrification is a universal trait in beta-proteobacterial ammonia oxidizers (Shaw et al., 2006) and potentially a significant source of N2O production in soils as AOB dominates ammonia oxidation activity in most soils. In pure culture studies, Poth and Focht (1985) found that N2O production from nitrifier denitrification in cultures of Nitrosomonas europaea only happened under conditions of O2 stress. As O2 concentration declined, reduction of NO2 to N2O by nitrifiers occurred at increasing rates (Goreau et al., 1980; Lipschultz et al., 1981). In soils, therefore, nitrifier denitrification may contribute significantly to N2O production under suboxic conditions

136


(Wrage et al., 2005; Venterea, 2007). However, differentiating between nitrifier denitrification and other N2O-producing processes (e.g., nitrifier nitrification and nitrification-coupled denitrification) has been methodologically challenging. Acetylene (C2H2) at concentrations of 0.1e10 Pa (0.0001%e0.01% by volume) inhibits ammonia oxidation by autotrophs such as N. europaea (Berg et al., 1982; Hynes and Knowles, 1982) and therefore inhibits nitrifier nitrification, nitrifier denitrification, and nitrification-coupled denitrification. However, these concentrations of C2H2 do not inhibit ammonia oxidation by heterotrophs (Hynes and Knowles, 1982), which might be the source of N2O under certain conditions such as low pH, high O2 availability, and ample carbon substrates (Anderson et al., 1993). Robertson and Tiedje (1987) used 10 Pa C2H2 to inhibit N2O production by nitrifiers and 100 kPa O2 to inhibit N2O production by denitrifiers. They came to the conclusion that nitrifier denitrification was not an important N2O production pathway in the soil studied. In retrospect, however, the method used in this study was not able to support this conclusion because 10 Pa C2H2 also inhibits nitrifier denitrification, as mentioned earlier. Webster and Hopkins (1996) changed Robertson and Tiedje (1987)’s inhibition experiment slightly by adding one more treatment with both 10 Pa C2H2 and 100 kPa O2 and concluded that nitrifier denitrification was the main source of N2O from the drier sandy loam soil. It was acknowledged by both Robertson and Tiedje (1987) and Webster and Hopkins (1996) that poor diffusion of the gases into the soil, especially the wet soil, limited the effectiveness of the inhibition method. Most importantly, the negative effect of 100 kPa O2 on ammonia oxidation itself likely underestimates the contribution of nitrifier denitrification, although 100 kPa O2 seemed to inhibit N2O production from nitrifier denitrification to a large extent (Wrage et al., 2004b). In grassland soils, Wrage et al. (2004a) used the same method as Webster and Hopkins (1996) to assess sources of N2O production and further revealed the problems related to the use of the inhibitors C2H2 and O2 to investigate nitrifier denitrification. For example, nearly one-third of all incubations with inhibitors produced more N2O than the controls, leading to the conclusion that decreasing O2 content also decreased N2O production by nitrifier denitrification. After these problems had been revealed, Wrage et al. (2005) conceived a dual isotope (15N and 18O) approach to distinguish the pathway of nitrifier denitrification from other N2O-forming processes. Kool et al. (2009a,b) further refined this method by including 18O-labeled NO3 to quantify the exchange of oxygen atoms (O) between H2O and N oxides during denitrification and nitrification in soils. By using this method, Kool et al. (2010) and Kool et al. (2011) demonstrated definitive N2O production by nitrifier denitrification. They showed that the relative importance of heterotrophic denitrification and nitrifier denitrification in soil N2O production is a function of water content. For example, when soil moisture conditions are suboptimal for heterotrophic denitrification in a sandy soil, nitrifier denitrification is the major contributor to N2O emission. By using this same dual isotope method, other environmental factors such as O2 availability and fertilizer N source were also found to control N2O


137

production through nitrifier denitrification (Zhu et al., 2013a). Although this method can differentiate pathways and sources of N2O production, its broad adoption may be limited by substantial assumptions and complicated calculations (Kool et al., 2010). It was reported that the 15N fractionation factor associated with nitrifier nitrification differs from heterotrophic denitrification, and the site preference (difference in d15N values between the two N atoms in N2O) following NH2OH decomposition differs from that of heterotrophic denitrification (Sutka et al., 2006; Frame and Casciotti, 2010). As a result, both the 15N/14N and 18 O/16O ratios of N2O in natural abundance measurements and site preference of N2O from nitrifier nitrification can differ from nitrification-coupled denitrification. This difference, in combination with the use of enriched 15N as a tracer, could permit direct quantification of each N2O production pathway. Although N2O production by nitrifier denitrification has been identified and quantified, knowledge gaps still exist about this pathway, including its ecological relevance and the physiological importance of NO2 as a substrate for AOB to produce N2O. Ritchie and Nicholas (1972) mentioned the possibility of NO2 as an alternative terminal electron acceptor to O2 under conditions of temporary anaerobiosis. In addition, Poth and Focht (1985) proposed three possible reasons for the use of NO2 as a terminal electron acceptor by autotrophic nitrifiers under limited O2 conditions: (1) conserving O2 for the initial mixed-function oxidase step of ammonium oxidation; (2) removing NO2 as a toxic product; and (3) decreasing competition for O2 by consuming the substrate for NO2 oxidizers. A fourth reason is possible: autotrophic nitrifiers gain energy through the reduction of NO2 when nitrifiers are not able to acquire enough energy from ammonia oxidation, which requires O2 and can cease under anoxic conditions (Hollocher et al., 1981; Andersson and Hooper, 1983).

FIGURE 6.3 Land management practices and factors affecting ammonia oxidation pathways in soils. Blue (gray in print version) arrows represent the relationships between soil factors and ammonia oxidation pathways; black arrows represent the relationships between land management practices and soil factors that can be directly changed by these practices.

138


After NO2 is produced in the nitrifier cell, it is unlikely to be transported outside of the cell to be utilized by nitrifiers again (Poth and Focht, 1985). This conclusion is based on experiments in which the addition of NO2 under conditions of O2 stress was not able to promote N2O production by Nitrosomonas europaea (Poth and Focht (1985). The reduction of NO2 must therefore occur within the nitrifier cell. However, the mechanism by which nitrifiers retain NO2 in the cell instead of transporting it out of the cell is unknown. The nitrifier cell may retain NO2 because the cell membrane cannot produce enough energy to actively transport NO2 out of the cell under limited O2 conditions. Consequently, it must be consumed to gain energy and to avoid its toxicity. This hypothesis is based on observations that nitrifier denitrification only occurs under conditions of O2 stress (Poth and Focht, 1985), and N2O produced by this pathway occurs at increasing rates as O2 concentration declines (Goreau et al., 1980; Lipschultz et al., 1981). Future studies are needed to test this hypothesis and address the critical gap in knowledge of the existence and importance of nitrifier denitrification.

Nitrification-Coupled Denitrification Differing from nitrifier denitrification, nitrification-coupled denitrification is the process by which NO2 or NO3 produced during nitrification is utilized by denitrifiers (Wrage et al., 2001). Nitrification-coupled denitrification is composed of two linked processes that produce N2O as an intermediate in the denitrification process. This pathway usually occurs in soils where favorable conditions for both nitrification and denitrification are present, such as in distinct neighboring microhabitats (Arah, 1997). Nitrous oxide production from this pathway is mainly produced at oxiceanoxic interfaces, likely when substrates diffuse to the surface of soil aggregates (Khdyer and Cho, 1983). For example, nitrification usually takes place in the oxic surface layers or cracks of the soil matrix, while denitrification mostly exists in the anoxic deeper layers and waterlogged areas and within aggregates (Tiedje et al., 1984; Leffelaar, 1986). The interface between where these processes occur is the area where N2O production through nitrification-coupled denitrification is highest. This process is also commonly employed in wastewater treatment to promote high removal rates of N by providing conditions that stimulate the linkage between nitrification and denitrification (dos Santos et al., 1996). In soil, Zhu et al. (2013a) observed that the relative importance of nitrification-coupled denitrification for total N2O emissions decreased as O2 concentration decreased. This was attributed to control of the overall rate of N2O production by the nitrifying step, which was depressed by the lower O2 concentration.

Factors Influencing N2O Production in Soils via Ammonia Oxidation Pathways Soil nitrifier activity, pH, ammonium ðNH4 þ Þ concentration, soil texture, temperature, moisture, O2 availability, and bioavailability of iron (Fe) and


139

manganese (Mn) can all influence N2O production from ammonia oxidation pathways (Fig. 6.3). Nitrous oxide production has been reported to be greater in clay soils than in loam or sandy soils (Zhu et al., 2013a), likely related to greater soil nitrifier populations and nitrification rates in finely textured soils (Fortuna et al., 2012). Nitrifier populations inhabiting clay surfaces have been shown to be protected from the effects of Hþ produced from ammonia oxidation (Powell and Prosser, 1991), thus favoring higher nitrifier populations and nitrification rates in clay soils compared to other soils. Also, short-term increases in soil pH limit NO2 oxidation mediated by Nitrobacter spp., resulting in the accumulation of NO2 (Hawkins et al., 2010) and contributing to N2O production (Venterea and Rolston, 2000; Venterea, 2007). In addition, soil pH plays a key role in shaping the communities of ammonia oxidizers that control N2O production processes (Gubry-Rangin et al., 2011). For example, AOA have more important role than AOB in ammonia oxidation process in acidic soils (Zhang et al., 2012), while AOB is the dominant group of ammonia oxidizers in alkaline soils (Jiang et al., 2015). In a physiological study, soil AOA strain Nitrososphaera viennensis and marine AOA strain Nitrososphaera maritimus have been reported to produce N2O in the same range as those of the AOB Nitrosospira multiformis; however, N2O production of AOA is not affected by the oxygen concentration (Stieglmeier et al., 2014). Generally, increases in temperature increase N2O production through ammonia oxidation pathways, especially nitrifier nitrification, by increasing ammonia oxidizer populations (Avrhami and Conrad, 2003; Avrahami and Bohannan, 2007; Szukics et al., 2010) and activity (Avrahami et al., 2003). However, Avrahami et al. (2003) found that when soil temperature increased from 25 to 37 C, potential nitrification decreased but the rate of N2O production increased monotonically. This indicates that under certain conditions, N2O production through ammonia oxidation pathways can be independent from nitrification. Oxygen availability in soil has been reported to be an important factor influencing N2O production through ammonia oxidation pathways, especially nitrifier denitrification (Zhu et al., 2013a). Nitrous oxide production via ammonia oxidation pathways has been observed to increase as O2 concentrations decreased from 21% to 0.5%, in agreement with results obtained from pure culture studies (Goreau et al., 1980; Lipschultz et al., 1981). Soil moisture influences N2O production by controlling not only the diffusion of O2 but also substrate availability (Stark and Firestone, 1995) and microbial activity (Avrahami and Bohannan, 2007). Metals, notably Fe, are essential cofactors for the transfer of electrons in many enzymes participating in ammonia oxidation pathways, such as ammonia monooxygenase, hydroxylamine oxidoreductase, and nitrite oxidoreductase (Fig. 6.2) (Meiklejohn, 1952; Godfrey and Glass, 2011; Stein, 2011; Glass and Orphan, 2012). The biological availability of Fe is a key factor that influences Fe-dependent enzymes. For example, cells of Nitrosomonas europaea that grow in Fe-limited medium have low heme and cellular Fe contents, which lead to decreases in the specific activities of ammonia monooxygenase and

140


FIGURE 6.4 Land management practices and factors affecting heterotrophic denitrification in soils. Blue (gray in print version) arrows represent the relationships between soil factors and heterotrophic denitrification; black arrows represent the relationships between land management practices and soil factors that can be directly changed by these practices.

hydroxylamine oxidoreductase compared to Fe-replete cells (Wei et al., 2006). This results in reductions in the ammonia oxidation rate and, thus, N2O production. Abiotic reactions between redox-active Fe and enzymatically derived reactive intermediates in ammonia oxidation (e.g., hydroxylamine) can also be a significant source of N2O (Zhu-Barker et al., 2015a). Details of such reactions are described in the section on Abiotic N2O production from soils.

HETEROTROPHIC DENITRIFICATION Heterotrophic denitrification is an anaerobic respiratory process performed by heterotrophic denitrifiers using NO3 or NO2 in place of O2 as alternate electron acceptors to consume organic matter (Fig. 6.2) (Firestone et al., 1980; Knowles, 1982). It is a stepwise reduction of NO3 or NO2 to N2 as the end product. Some intermediates, such as NO and N2O, can be released into the atmosphere depending on environmental conditions (Firestone et al., 1979, 1980; Firestone and Davidson, 1989a; Weier et al., 1993; Mulvaney et al., 1997; Van Cleemput, 1998; Gillam et al., 2008; Phillips, 2008). These reduction reactions are carried out by denitrifiers representing a broad range of bacteria taxa, including Pseudomonas, Bacillus, Thiobacillus, Propionibacterium, and others (Firestone, 1982). Enzymes catalyzing the reductions are nitrate reductase, nitrite reductase, nitric oxide reductase, and nitrous oxide reductase (Hochstein and Tomlinson, 1988).

Factors Influence N2O Production Through Heterotrophic Denitrification in Soils While many factors influence the denitrification process, water content is commonly identified as the most important factor in soil through its control

Heterotrophic Denitrification

141

on O2 diffusion, C availability, and NO3 concentration (Fig. 6.4) (Smid and Beauchamp, 1976; Firestone et al., 1979; Firestone and Davidson, 1989a; Weier et al., 1993). Findings from multiple laboratory incubations at constant temperature and field studies indicate that denitrification-derived N2O increases as soil water content increases, especially when soil water content is higher than 70% of water-filled pore space (Smith and Tiedje, 1979; Myrold and Tiedje, 1985; Grundmann et al., 1988; Schuster and Conrad, 1992; Weier et al., 1993; De Klein and Van Logtestijn, 1994; Maag and Vinther, 1996; Godde and Conrad, 2000; Bateman and Baggs, 2005; Menyailo and Hungate, 2006). However, once soil water content reaches saturation or flooding, N2O production by heterotrophic denitrification decrease precipitously because a strict anoxic environment often leads to complete denitrification (Firestone et al., 1979; Bonin et al., 1989). Oxygen concentration has distinct effects on the different enzymes associated with denitrification. Nitrate reductase is less sensitive to O2 than NO2 and N2O reductases. Nitrate reductase can be completely inhibited only at O2 concentrations greater than 0.25%, compared with 0.13% and 0.02% for NO2 and N2O reductases, respectively (Bonin et al., 1989). Thus, the amount of N2O released during denitrification is closely related to O2 concentration in the soil environment (Hochstein et al., 1984). Cooper and Smith (1963) reported that the overall rate of denitrification did not change when the NO2 concentration increased from 37.5 to 150 ppm, but increases in NO3 concentration resulted in a higher proportion of N2O as the end product (Cooper and Smith, 1963; Firestone et al., 1979; Knowles, 1982; Weier et al., 1993), likely because NO3 is a preferred electron acceptor over N2O (Thauer et al., 1977; Firestone et al., 1980; Firestone and Davidson, 1989b). It was reported that N2O consumption is limited when NO3 concentration is higher than 5e10 mg N/kg soil (Ryden, 1983; Gillam et al., 2008). The amount of N2O that is released during denitrification is also higher if the pH is low because N2O reductase is inhibited at low pH (Knowles, 1982). Considering the multiple effects of NO3 and pH on N2O derived from denitrification, the use of an average N2O/N2 ratio for estimation of denitrification from N2O field measurements is not recommended (Weier et al., 1993). Other than O2 availability and NO3 concentration, soil organic C as an energy source and electron donor for denitrifiers is also highly correlated with denitrification potential (Myrold and Tiedje, 1985; Weier et al., 1993; Gillam et al., 2008). In other words, both N2O and N2/N2O ratio increase as soil organic C, especially decomposable C, increase (Burford and Bremner, 1975). Temperature is another important factor that influences denitrification. For example, the rate of N2O þ N2 evolved from denitrification was shown to increase when soil temperature increased to 67 C, but N2O became an increasingly smaller component of the gaseous N evolved, persisting for shorter times with increasing temperature (No¨mmik, 1956; Smid and Beauchamp, 1976; Keeney et al., 1979). Also, addition of copper, a cofactor of N2O reductase, has been found to promote the reduction of N2O to N2 (Rasmussen et al., 2000; Paraskevopoulos et al., 2006; Wang et al., 2013).

142


OTHER BIOLOGICAL PROCESSES Apart from heterotrophic denitrification and ammonia oxidation pathways, other biological processes, such as heterotrophic nitrification, fungal denitrification, DNRA, feammox, and comammox, also contribute to N2O production in soils under certain conditions.

Heterotrophic Nitrification In contrast to ammonia oxidation, in which autotrophic nitrifiers use CO2 as a C source, heterotrophic nitrifiers use organic C as a C source (Robertson and Kuenen, 1990). Many studies have reported that heterotrophic nitrifiers can use both ammonia and organic N compounds as substrates (Papen et al., 1989; Barraclough, 1995; Honda et al., 1998; De Boer and Kowalchuk, 2001; Islam et al., 2007; Zhang et al., 2014). Although the intermediates and products of heterotrophic and autotrophic nitrification are the same, the enzymes involved in these two pathways differ. For example, ammonia monooxygenase from the heterotrophic nitrifier Pseudomonas denitrificans cannot be inhibited by C2H2. Also, the hydroxylamine oxidoreductase present in heterotrophic nitrifiers is a nonheme Fe enzyme, unlike the multiheme enzyme found in autotrophic nitrifiers (Richardson et al., 1998; Braker and Conrad, 2011). As many heterotrophic nitrifiers are able to denitrify under aerobic conditions, N2O is assumed to result as an intermediate during the reduction of NO2 to N2 in this pathway (Robertson et al., 1989; Anderson et al., 1993). In pure culture studies, heterotrophic nitrifiers produce more N2O per cell than autotrophic nitrifiers (Papen et al., 1989; Anderson et al., 1993). Blagodatsky et al. (2006) observed that the maximum rate of N2O production occurred after a sudden decrease of O2 in the medium. However, when the O2 concentration dropped lower than 2%, heterotrophic nitrification ceased altogether (Anderson et al., 1993). Heterotrophic nitrification is prevalent in acidic soils and plays a predominant role in NO3 and N2O production in these soils (Schimel et al., 1984; Kreitinger et al., 1985; Stroo et al., 1986; Wood, 1990; Pedersen et al., 1999; Huygens et al., 2008; Zhang et al., 2011, 2014). In contrast, the activity of autotrophic nitrifiers is very low in a low pH environment (Weber and Gainey, 1962). However, the importance of heterotrophic nitrification varies in soils under different land uses; it has been shown to be responsible for about 18%, 67%, 78%, and 92% of total nitrification, respectively, in forest clear-cuts, the organic horizon of a mature forest, a young forest, and the mineral soil of a mature forest (Pedersen et al., 1999). The contribution of heterotrophic nitrification to overall N2O production is therefore significant under soil conditions such as low pH, high O2 concentration environments, and high organic matter content.

Fungal Denitrification Fungal denitrification is a unique N2O production process mediated by fungi (eukaryotes) (Bollag and Tung, 1972; Burth and Ottow, 1983; Shoun et al., 1992).

Other Biological Processes

143

As fungi possess the ability to simultaneously perform both denitrification and O2 respiration, this pathway is also called aerobic denitrification (Robertson and Kuenen, 1990; Zhou et al., 2001). It has been suggested that fungi lack N2O reductase, which indicates that this process may be a significant source of N2O production in soils dominated by fungi such as those in forests and grasslands (Laughlin and Stevens, 2002). Interestingly, fungal denitrification is often accompanied by codenitrification, in which a hybrid N2 or N2O is formed when one N atom from the NO3 pool combines with one N atom from N donors such as amines and imines. Readers are referred to Shoun et al. (2012) for a detailed review of fungal denitrification systems.

Dissimilatory NO3 Reduction to NH4 þ In anoxic conditions, DNRA is an alternative reduction pathway that can be catalyzed by numerous bacterial genera (Cole and Brown, 1980; Cole, 1988; Stevens et al., 1998). Nitrous oxide can be produced during both DNRA and heterotrophic denitrification, while N2 is produced only by heterotrophic denitrification. In the DNRA pathway, one set of fermentative bacteria is responsible for formation of NO2 (Cole and Brown, 1980) while another set of fermentative bacteria complete this pathway by reducing NO2 to NH4 þ, during which N2O may be released as well. No N2O reductase has been found in these bacteria (Smith, 1982). Tiedje et al. (1983, 1988) reported that compared to heterotrophic denitrification, which prefers a medium with a low C:N ratio, DNRA is more likely to occur in environments abundant in C but low in available N. In environments with ample NO3 , competition for C will determine which NO3 reduction process is favored (Tiedje, 1988). Some studies have suggested that DNRA can be a significant source of N2O production in soils (Stanford et al., 1975; Caskey and Tiedje, 1979; Silver et al., 2001;

FIGURE 6.5 Schematic representation of the abiotic processes in which iron, manganese, and organic matter interact with nitrogen species in soils. Adapted from Zhu-Barker, X., Horwath, W.R., Burger, M., 2015c. Knife-injected anhydrous ammonia increases yield-scaled N2O emissions compared to broadcast or band-applied ammonium sulfate in wheat. Agriculture, Ecosystems and Environment 212, 148e157.

144


Ru¨tting et al., 2011). In tropical forest soils, Silver et al. (2001) observed that DNRA accounted for approximately 75% of the turnover of the NO3 pool and that in these systems, the rate of DNRA was primarily limited by NO3 availability, rather than C or O2. To date, no methodology has been developed to differentiate N2O production between DNRA and heterotrophic denitrification.

Feammox The reduction of Fe(III) in soil can be coupled to anaerobic NH4 þ oxidation in a processes termed feammox. Nitrate, NO2 , and N2 are confirmed products in this pathway (Cle´ment et al., 2005; Shrestha et al., 2009; Yang et al., 2012). However, production of N2O has also been observed (Li et al., 1988). In the 1980s, Li et al. (1988) first found that Fe(III) acts as an electron acceptor for anaerobic NH4 þ oxidation in two paddy soils. Owing to the lack of available isotopic methods at the time, they did not quantify the amount of N loss through this pathway. By incubating Fe-rich riparian soils under strictly anoxic conditions with no initial NO3 or manganese, Cle´ment et al. (2005) observed the unexpected production of NO2 and Fe(II). They then proposed that a biological process can use Fe(III) as an electron acceptor while oxidizing NH4 þ to NO2 for energy production. In the same field where the soil was collected for Cle´ment et al. (2005)’s study, Shrestha et al. (2009) detected the production of NO2 in situ when no O2 and initial NO3 were present in the soil environment. These findings indicate that feammox as an alternative N loss pathway indeed exists in soil. Yang et al. (2012) reported that rates of feammox can be comparable to aerobic nitrification and denitrification. Thus, feammox may drive N losses in soils rich in poorly crystalline Fe oxide minerals and under low or fluctuating redox conditions. However, the extent to which feammox contributes to N2O production in soils is still unknown, as is the mechanism of this pathway.

Comammox Nitrification has traditionally been considered to be a two-step process catalyzed by chemolithoautotrophic microorganisms oxidizing either NH3 or NO2 (Winogradsky, 1890). In contrast, comammox is the complete oxidation of NH3 to NO3 by one organism (Daims et al., 2015; van Kessel et al., 2015). This process is energetically feasible in certain nitrifiers. So far, two Nitrospira species have been characterized for the ability to perform comammox (van Kessel et al., 2015). Ammonia monooxygenase and NH2OH dehydrogenase genes for ammonia oxidation, and NO2 oxidoreductase for NO2 oxidation, have been found in these species. It allows for comammox, like nitrifier nitrification, to be a potential source of N2O. Nevertheless, the lack of assimilatory NO2 reductase in the studied Nitrospira species indicates that any production of N2O by comammox is unlikely to result from the reduction of NO2 in the cell. However, the two Nitrospira species observed by van Kessel et al. (2015) are not the only organisms in the environment known to conduct comammox;

Hydroxylamine Decomposition

145

other species containing NO2 reductase may in fact exist in the comammox community and contribute to N2O production. This novel pathway has recently been identified in wastewater systems and ocean ecosystems (Chao et al., 2016; Pinto et al., 2016), but it is also believed to exist in soils. Future studies are needed to understand its ecological function and contribution to global N cycling and N2O production.

ABIOTIC N2O PRODUCTION IN SOILS Occurrence of abiotic reactions that produce N2O has been known for nearly a century (Bray et al., 1919). Iron (Fe), Mn, and organic compounds readily engage in redox reactions with intermediates in the N cycle, leading to production of N2O under relevant environmental conditions (Fig. 6.5). Although biological processes are commonly assumed to be the major source of N2O production in soil, abiotic reactions between these redox-active metals, organic matter, and N intermediates can be significant sources as well. This section summarizes abiotic pathways by which N2O can be produced in soils. For details on defining characteristics, environmental controls, and isotopic aspects of abiotic reactions that lead to N2O production, readers are referred to the review of Zhu-Barker et al. (2015a).

HYDROXYLAMINE DECOMPOSITION Production of N2O from the chemical oxidation of hydroxylamine by Mn(IV) and Fe(III) oxides, typified in Eqs. (6.1) and (6.2), has been assumed to occur in soil since the mid-20th century (Mann and Quastel, 1946; Bremner and Shaw, 1958). Evidence for this reaction was first reported by Nelson and Bremner (1970), who observed significant amounts of N2O produced in soils through the reaction of Fe(III) with hydroxylamine. In a follow-up study, N2O yield by chemical decomposition of hydroxylamine in diverse soil types was found to be faster and greater than N2 production (Bremner et al., 1980). In this study, production of N2O was highly correlated with exchangeable and oxidized Mn(IV) in the soils studied, and N2O generated via hydroxylamine decomposition greatly exceeded N2O generated via chemodenitrification (see the Chemodenitrification section) or the reaction of NO2 with hydroxylamine. Similar results showing that NO2 additions did not increase hydroxylamine oxidation to N2O were also reported in another study (Minami and Fukushi, 1986). 4 a-FeOOH þ 2NH2OH þ 8Hþ / N2O þ 4Fe(II) þ 9H2O DG0 ¼ 6.3 kJ mol/e (6.1) DG0

2MnO2 þ 2NH2OH þ 4Hþ / N2O þ 2Mn(II) þ 5H2O ¼ 106.2 kJ mol/e (6.2)

Despite strong evidence for the importance of abiotic hydroxylamine decomposition in soil N2O emissions, Bremner (1997) concluded that this process was unimportant for N2O production in soil, based on the lack of evidence

146


for extracellular release of hydroxylamine during nitrification. Since this report, most soil research has focused on biological/enzymatic N2O production, leaving N2O emission from abiotic NH2OH decomposition processes largely ignored. However, recent interest in “cryptic” elemental cycling has reinvigorated the study of highly reactive N species. Recently, Liu et al. (2014) developed a novel, highly sensitive method to quantify concentrations of hydroxylamine in soil ranging from 0.3 to 34.8 mg N/kg soil. Significant N2O was produced when cropland soils were amended with NH2OH, but not forest soils, which can have lower pH (Heil et al., 2015). Soil pH, Mn content, and C: N ratio can significantly affect hydroxylamine-induced N2O emissions. For example, more hydroxylamine-induced N2O production was observed at higher soil pH and higher Mn content, while a negative correlation was observed for the relationship between hydroxylamine-induced N2O emissions and C: N ratio (Heil et al., 2015).

CHEMODENITRIFICATION Chemodenitrification is the chemical decomposition of N intermediates in ammonia oxidation (i.e., NO2 ) or denitrification (i.e., NO3 , NO2 , NO, or N2O) coupled with the oxidation of Fe(II) (e.g., Eqs. 6.3e6.6) (Chao and Kroontje, 1966; Buresh and Moraghan, 1976; Chalk and Smith, 1983; Sørensen and Thorling, 1991; Van Cleemput, 1998; Picardal, 2012). Soil N2O production from chemodenitrification is stimulated by decreasing O2 concentration, increasing soluble Fe(II), and the presence of organic matter (Nelson and Bremner, 1970; Van Cleemput and Baert, 1984). Therefore, N2O production from chemodenitrification is assumed to operate mostly under anoxic conditions with a high supply of organic matter, such as in wetlands and waterlogged soils. In these environments, persistence of significant concentrations of Fe(II) and bacterial denitrification can lead to the accumulation of reactive NO2 and NO (Van Cleemput, 1998). Under anoxic conditions and at circumneutral pH, humic acids reduce NO to N2O (Stevenson et al., 1970) while diverse organic molecules (e.g., fulvic acid, lignins, and phenols) reduce nitrous acid to N2 and N2O (Stevenson and Swaby, 1964). In aerobic soils, significant abiotic N2O production from NO2 decomposition (31%e75% of total emissions) coupled to other electron donors such as humic acids also exists (Thorn and Mikita, 2000; Venterea, 2007). DG0

H

þ

NO3 þ 2Fe(II) þ 3H2O / NO2 þ 2 a-FeOOH þ 4Hþ ¼ 96.8 kJ mol/e (6.3)

DG0

DG0

¼ 88.7 kJ mol/e

NO2 þ Fe(II) þ H2O / NO þ a-FeOOH þ (6.4)

2NO þ 2Fe(II) þ 3H2O / N2O þ 2 a-FeOOH þ Hþ ¼ 168.3 kJ mol/e (6.5)

Fertilization

Hþ DG0 ¼ 316.7 kJ mol/e

147

N2O þ Fe(II) þ H2O / N2 þ a-FeOOH þ (6.6)

Recent studies of Antarctic hypersaline lakes have also reinvigorated interest in chemodenitrification with Fe(II). Samarkin et al. (2010) documented large fluxes of N2O from soils taken near Don Juan Pond, a metabolically dormant hypersaline pond in southern Victoria Land, Antarctica. The authors concluded that chemodenitrification was the major source of N2O due to the apparent lack of biological activity and high concentrations of NO3 , NO2 , and Fe(II) minerals in the brine. However, isotopic analysis of this putatively abiotic N2O did not reveal any clear distinguishing differences from biotic isotopic signatures (Peters et al., 2014).

LAND MANAGEMENT PRACTICES TO CONTROL N2O EMISSION FROM SOILS Cropping systems are sources and consumers of N2O and CH4 emissions and there is a large potential for changes in management to mitigate these GHG emissions (Smith et al., 2008; USDA, 2014). In fact, N2O emissions that correspond to different agricultural and soil management practices are a key focus of GHG emission concerns in upland cropping systems (USEPA, 2016). GHG emissions from agriculture are highly variable in time and space due to numerous factors such as water content, O2 levels, pH, and the availability of C and N (Bremner et al., 1980; Bouwman et al., 2002b; Bateman and Baggs, 2005; Stehfest and Bouwman, 2006; Zhu et al., 2013a). Management practices influence N2O emissions from soils by altering these edaphic factors and the activities of microorganisms involved in N2O production (Figs. 6.3 and 6.4) (Avrahami et al., 2002; Avrahami and Bohannan, 2007; Garland et al., 2011; Venterea et al., 2011; Kennedy et al., 2013; Zhu-Barker et al., 2015b,c). This section provides a summary of current knowledge and focuses on how management practices influence N2O emissions in upland cropping systems, as well as which practices could be employed to mitigate N2O emissions from agriculture.

FERTILIZATION Nitrogen is generally the most important nutrient for crop growth and an essential input to maintain high crop yields. The management of synthetic and organic N additions to cropland soils is the main determinant of N2O emissions in the agricultural sector (Bouwman et al., 2002a; Davidson, 2009; USEPA, 2016). Once fertilizer N enters into soil it can be directly converted to N2O by soil microorganisms and by abiotic reactions discussed in the preceding sections of this chapter. Other forms of N which are formed as a result of fertilizer application, e.g., ammonia, NO, and NO3 , may also be converted to N2O; such emissions are referred as “indirect” N2O emissions (de Klein et al., 2006; Beaulieu et al., 2011). Fertilizers drive N2O emissions from soil through their influence on the availability of N substrates for soil microorganisms

148


(e.g., nitrifiers and denitrifiers) and abiotic reactions, and through their influence on soil pH and O2 availability (Harrison et al., 1995; Mulvaney et al., 1997; Smith et al., 1998; Venterea and Rolston, 2000; Zhu et al., 2014b). In the subsequent subsections, the effects of fertilizer rate, placement, timing of application, and efficiency enhancers on N2O emissions will be discussed.

Fertilizer Rate Evidence from numerous studies indicates that the application rate of N fertilizer strongly influences N2O emissions (Grant et al., 2006; Zebarth et al., 2008b; Burger and Horwath, 2012; Zhu et al., 2014b). Applied N fertilizers provide substrates for biotic and abiotic processes of the N cycle, thus promoting N2O production from soil. Generally, a linear relationship between N application rate and N2O emissions is assumed, and so this permits use of a default emission factor by the Intergovernmental Panel on Climate Change (IPCC) (IPCC, 2007b). Most national GHG inventories have been constructed using this emission factor approach (de Klein et al., 2006; USDA, 2014). The N2O emission factor is defined as the percentage of fertilizer N applied that is emitted as N2O. It is calculated as the difference in emissions between fertilized and unfertilized soil under otherwise identical conditions. The current global mean emission factor for fertilizer-induced direct N2O emission is w0.9% (Bouwman et al., 2002b; Stehfest and Bouwman, 2006), which is an approximate average of emissions induced by synthetic fertilizer (1.0%) and manure (0.8%), and it has been rounded up by the IPCC to 1% to account for uncertainties (de Klein et al., 2006). For example, for every 100 and 200 kg of N input, 1.0 and 2.0 kg N are emitted directly from soil as N2O, respectively. However, a growing number of field studies with multiple N fertilizer rates indicate that N2O emissions in fact respond nonlinearly to increasing N rates across a range of fertilizer types, climate, and soil textures (McSwiney and Robertson, 2005; Ma et al., 2010; Hoben et al., 2011; Zhu-Barker et al., 2015c). Using a meta-analysis of data collected from 78 published studies with 233 site-years and at least three N input levels, Shcherbak (2014) also substantiated a nonlinear response of soil N2O emissions to fertilizer N rates. It has long been recognized that free ammonia acts as a NO2 oxidation inhibitor (Anthonisen et al., 1976; Turk and Mavinic, 1989; Yun and Kim, 2003; Simm et al., 2006). Thus, an increase in the application rate of ammoniacal N fertilizer increases the potential for NO2 accumulation and associated N2O production (Van Cleemput and Samater, 1996; Venterea, 2007). As described previously, O2 availability is a key factor controlling nitrification rate and N2O production through ammonia oxidation (Goreau et al., 1980; Bollmann and Conrad, 1998; Zhu et al., 2013a). The process of nitrification is a process consuming O2, which acts as an electron acceptor during the ammonia oxidation and NO2 oxidation steps (Lees and Simpson, 1957; Butt and Lees, 1960; Hollocher et al., 1981). Increases in ammoniacal N fertilizer application promote nitrification which increases O2 consumption in the environment

Fertilization

149

(Zhu et al., 2014b). This in turn stimulates N2O production through nitrifier denitrification and heterotrophic denitrification (Zhu et al., 2013a). Therefore, the Tier one N2O accounting method adopted by IPCC, whereby N2O emissions are assumed to be a simple fraction of N inputs, potentially underestimates or overestimates fertilizer-derived N2O emissions. Regardless of other fertilization practices such as fertilizer type, placement, and application timing (see sections on fertilizer type, fertilizer placement, and fertilizer application timing), the rate of N fertilizer application itself can be refined to reduce N2O emissions as long as rates are not reduced to the point that yields decline. Reducing N rates below what is required to support optimal crop growth can also promote organic matter mineralization and increase residual effects of fertilizer, leading to soil C loss and an overall increase in GHG emissions. Yield-scaled emissions have recently emerged as a means to recognize the highly valued provisioning services of agriculture with respect to associated GHG emissions when discussing agricultural sources of N2O (Van Groenigen et al., 2010; Murray and Baker, 2011; Zhu-Barker et al., 2015c). Evaluating crop yield per unit, N2O emissions balances the inherent tradeoff between food productivity and N2O emissions and reflects the efficiency of N fertilizer use in the system (Van Groenigen et al., 2010; Venterea et al., 2011; Zhu-Barker et al., 2015c). For example, Zhu-Barker et al. (2015c) reported that in a wheat cropping system, yield-scaled N2O decreased by 59% when N fertilizer rate was reduced from 266 to 154 kg N/ha, while yields were not significantly affected by fertilizer rate. In a lettuce field, however, yieldscaled N2O emissions were not affected when N fertilizer rate was reduced from 168 to 84 kg N/ha (Burger and Horwath, 2012). In general, practices that increase crop N use efficiency (NUE) are expected to reduce N2O emissions because the applied N that is taken up by crops is not available to soil microorganisms that produce N2O. However, strategies that improve NUE cannot always reduce N2O emissions. Other practices, such as fertilizer types and placement, can result in varying N2O emissions irrespective of the effect on NUE (Gagnon and Ziadi, 2010; Gagnon et al., 2011; Zhu-Barker et al., 2015c).

Fertilizer Type The role of fertilizer type in generation of N2O includes N2O emissions associated with the manufacture and transportation of fertilizer and the direct emissions from soils following fertilizer application. This section focuses only on the direct N2O emissions from soil following the fertilizer application. For the details of GHG emissions associated with the production of different types of fertilizers, readers are referred to Snyder et al. (2009) and Burger and Venterea (2011). Nitrogen fertilizer type influences direct N2O emissions from soils mainly by determining the form of the N substrate and the short- and long-term changes in soil pH. Depending on the direction of the instant soil pH change after

150


application, these fertilizers can be divided into alkaline-forming (e.g., anhydrous ammonia and urea) and acidifying fertilizers (e.g., ammonium sulfate and ammonium nitrate). In the United States, the primary synthetic N fertilizers used are urea ammonium nitrate, calcium ammonium nitrate, anhydrous ammonia, ammonium sulfate, and urea. Respectively, these had annual average sales of 0.576, 0.258, 0.237, 0.177, and 0.150 Mg between 2007 and 2012 in the United States (CDFA, 2012; ERS, 2016). These fertilizers are ammonia-based and their long-term application generates soil acidity through nitrification (Hauck and Stephenson, 1965; Bouman et al., 1995; Mulvaney et al., 1997). Acidifying fertilizers like these tend to promote higher N2O emissions when denitrification under anoxic soil conditions predominates, especially in soils that have a low initial pH (Mulvaney et al., 1997). This is because at low pH, the substrate affinities of the enzymes NO3 reductase is higher than the N2O reductase (Betlach and Tiedje, 1981; Ottow et al., 1985; Na¨gele and Conrad, 1990). Nitrification rates are generally higher for alkaline-forming than for acidifying N fertilizers (Hauck and Stephenson, 1965; Mulvaney et al., 1997). Initial increases in soil pH after application of alkaline fertilizers can last several weeks and result in the accumulation of NO2 (Hawkins et al., 2010), which is a substrate for N2O production (Venterea and Rolston, 2000; Venterea, 2007). Under anoxic conditions, alkaline-forming fertilizers can also stimulate denitrification by increasing the solubility of soil organic matter (Norman et al., 1987). However, the effect of dissolved organic C on N2O emissions after applications of alkaline-forming fertilizers is difficult to ascertain because dissolved organic C can also promote the consumption of N2O and reduction to N2 (Burford and Bremner, 1975). A global review has led to the conclusion that the differences in emissions among fertilizer types are often marginal (Stehfest and Bouwman, 2006). This marginal difference emerges because the effect of N fertilizer type on N2O emissions varies across climates, crops, soil types, irrigation approaches, and fertilizer placement techniques. In side-by-side field trials, however, N2O emissions resulting from different types of fertilizers have been observed to differ significantly (Breitenbeck et al., 1980; Breitenbeck and Bremner, 1986; Thornton et al., 1996; Burton et al., 2008a; Halvorson et al., 2010a; Burger and Venterea, 2011; Zhu-Barker et al., 2015c). Studies in US corn cropping systems have found that anhydrous ammonia tend to increase N2O emissions compared to urea, urea ammonium nitrate, and other fertilizers. A field study of wheat cropping systems also demonstrated higher N2O emissions with anhydrous ammonia relative to urea under conventional tillage practice, although the opposite phenomenon was observed under no-till management (Burton et al., 2008a). Zhu-Barker et al. (2015c) reported that knife injection of anhydrous ammonia in wheat systems increased N2O emissions compared to use of ammonium sulfate. In particular, this practice increased yield-scaled N2O emissions compared to broadcast or band-applied ammonium sulfate. Fujinuma et al. (2011) also found 40% e200% higher N2O emissions from corn plots that received anhydrous ammonium than from plots that received urea. However, no difference between

Fertilization

151

anhydrous ammonia and urea was observed under certain crops (e.g., wheat) that received substantially lower N application rates than in the studies done with other crops (e.g., corn) (Burton et al., 2008a). Higher N2O emissions from urea applications compared to ammonium sulfate, ammonium nitrate, and calcium nitrate were also found in a series of field and laboratory studies (Tenuta and Beauchamp, 2003; Zhu et al., 2013a). In some of these experiments, both fertilizer type and placement varied among treatments because typical fertility practices were evaluated, and this may have influenced the results. In all the studies, anhydrous ammonia was injected while urea was either subsurface banded, broadcast, disked after broadcast, or applied through irrigation. In general, these studies tend to show that specific cropping systems, and presumably associated management, edaphic and climatic characteristics, contribute to variation in the magnitude of N2O emissions from different fertilizer types.

Fertilizer Placement Fertilizer placement can have a significant impact on both crop yield and N2O emissions. In the field, fertilizer can be applied through broadcasting, spraying, banding, subsurface banding, and fertigation (i.e., delivered at the surface, subsurface drip irrigation, knife injection, or microsprinklers), depending on the crop and the type of fertilizer. Improper placement of fertilizers can lead to the leakage of N into the environment causing diminished yield potential, decreased nutrient use efficiencies, and other related issues. The influence of fertilizer placement on N2O emissions is evident mainly through localized changes in N concentration, soil pH, and O2 availability. Generally, fertilizer band application results in a higher N concentration in the soil around the banding location compared to broadcast fertilizers. As stated earlier, concentrated ammoniacal fertilizers increase the potential for NO2 accumulation and associated N2O production (Mulvaney et al., 1997). In a greenhouse and field study in Montana, switching fertilizer placement from broadcasting to banding to nest application increased N2O production from 2.8 to 5.0e6.1 g N/ha d (Engel et al., 2010); this was due to the increased NO2 concentration, caused by an increase in urea and decrease in nitrification rate (Malhi et al., 1994). Similarly, in a field study in Colorado, Halvorson and Del Grosso (2013) found that surface-banded urea application led to higher N2O emissions than surface broadcast application when applied at 202 kg N ha1. However, no difference in N2O emissions was found between banded and broadcast applications when the application rate was relatively low (100 kg N/ha) (Engel et al., 2010). In Minnesota, Maharjan and Venterea (2013) also reported significantly higher N2O emissions from mid-row banding of 180 kg N/ha urea relative to broadcasting followed by incorporation. In contrast, lower N2O emissions from surface banding than from surface broadcasting were observed in a study conducted in Canada (Hultgreen, 2003), indicating the need to consider climate alongside fertilizer application techniques when scaling up field observations on the influence of fertilizer placement on N2O emissions.

152


The effect of fertilizer application depth on N2O emissions depends on tillage practices, which influence soil water content, O2 availability, and vertical distribution of C availability (Alcańtara et al., 2016). In a corn system in Eastern Canada, N2O emissions from shallow placement of ammonium nitrate were 26% lower than those from deep N placement (Drury et al., 2006), likely due to relatively lower O2 availability in deeper soil, while deep placement of liquid urea ammonium nitrate resulted in lower N2O emissions compared with shallow N placement (Liu et al., 2006). The highly volatile nature of anhydrous ammonia restricts its mode of application to subsurface injection, which generally results in a highly concentrated band of NH4 þ and consequent NO2 accumulation (Chalk et al., 1975; Venterea et al., 2010). In Iowa, Breitenbeck and Bremner (1986) reported that N2O emissions following injection of anhydrous ammonia at 30 cm were 107% and 21% greater than injections at 10 and 20 cm, respectively. However, lower seasonal N2O emissions were found following anhydrous ammonia applied at a depth of 15e20 cm than at a depth of 10e12 cm in irrigated sand used for corn production (Fujinuma et al., 2011). This discrepancy was likely caused by differences in the vertical distribution of C availability, which directly influences N2O consumption. Further studies will contribute to greater understanding of the interactions between fertilizer placement and soil types, climate, crops, and management regimes.

Fertilizer Application Timing Synchronizing N availability in the soil with crop demand for N is a major challenge in managing N fertilizer for crop production. Generally, crop N requirements are relatively low at seeding, increase several weeks after planting, and decrease sharply as the crop approaches maturity (Eagle et al., 2011). If N fertilizer is applied at the optimal stage, the active and well-developed root system can utilize fertilizer N most efficiently and reduce the potential for soil microbial and chemical processes to transform the applied N into N2O and other mobile forms such as NO3 , which can result in water pollution and contribute to indirect N2O emissions. Failure to adequately synchronize N fertilizer with crop requirements can cause an N surplus in soil. As the most commonly used fertilizer in US agriculture, excess N in the form of NH4 þ from ammoniacal fertilizer will promote nitrification, increasing the soil NO3 pool, and will increase N2O loss through both ammonia oxidation and heterotrophic denitrification once soil O2 is depleted (Zhu et al., 2013a). To demonstrate the effect of improved N application timing on N2O emissions, Burton et al. (2008b) compared a single application of 200 kg N/ha at planting versus a split application of 120 kg N/ha at planting plus 80 kg N/ha applied at final hilling in a potato cropping system. They found that emissions of N2O were reduced by 0.12e0.41 ton CO2e (CO2 equivalents) ha1 yr1 in the split application. Switching fertilizer application from fall to spring, from preplant to postplant, has also been shown to reduce N2O emissions in certain areas, i.e., areas with high rainfall or irrigation (Matson et al., 1998;

Fertilization

153

Hao et al., 2001; Hultgreen, 2003; Burton et al., 2008a). Phillips et al. (2009) reported that N2O emissions from dryland corn were 21% lower when N fertilizer was applied in mid-May as compared to early April. Zebarth et al. (2008a) observed that when N applications were at or in excess of corn crop N requirements, changing fertilizer application to side-dress, together with reducing the amount of N applied at each event, reduced the accumulation of NO3 but did not lead to any significant effect on cumulative N2O emissions. Optimizing the timing of N fertilizer application, therefore, may not always result in direct reductions of N2O emissions due to interactions with climate and conditions related to crop growth.

Fertilizer Efficiency Enhancers The development of controlled- or slow-release fertilizers and fertilizers containing inhibitors, collectively referred to as enhanced efficiency fertilizers, has been pursued in recent decades. Goals for these products include synchronizing N availability with crop requirement, improving NUE, stabilizing crop yields, and decreasing N losses through NO3 leaching and N2O emissions. These efficiency-enhancing products fall into three categories: (1) fertilizers coated with polymers, sulfur, and calcium magnesium phosphate, in which a physical barrier controls the release of plant available N; (2) nitrification inhibitors, which suppress nitrifier activity over a certain period of time and prevent the oxidation of NH4 þ to NO2 and consequently to NO3 ; and (3) urease inhibitors that slow down the rate of urea hydrolysis in the soil (Shaviv, 2001) and allow for a slow increase in soil NHþ 4 , favoring better synchrony between N availability and crop uptake over the course of the season (Trenkel, 2010). In recent years, application of enhanced efficiency fertilizers in field crop systems has received much attention. Several field studies conducted in various locations around the world have examined the possible effects of these products on N2O emissions (Oenema et al., 2001; Dalal et al., 2003; Akiyama et al., 2010). A meta-analysis assessing enhanced efficiency fertilizers and their effects on N2O emissions, using data from 20 field studies (Akiyama et al., 2010), concluded that the effects of polymer-coated fertilizers on N2O emissions are mixed and depend on specific site conditions. These products significantly reduced N2O emissions in imperfectly drained grasslands but were less effective in well-drained upland soils. Moreover, in a study conducted in loamy sand in Minnesota, Hyatt et al. (2010) reported a decrease in N2O emissions following the application of polymer-coated fertilizer compared with a conventional split application of 270 kg N/ha. However, Venterea et al. (2011) observed no impact or even higher emissions of N2O from plots receiving polymer-coated fertilizer. It is therefore difficult to generalize the effect of polymer-coated fertilizers on N2O emissions because the performance of these fertilizers varies widely across land use practices, soil types, and soil moisture regimes. However, they tend to be more effective when emission rates are higher and less effective when emissions rates are lower (Burger and Horwath, 2012).

154


Nitrification inhibitors can potentially decrease N2O emissions from ammonia oxidation and denitrification. In field studies, nitrification inhibitors have been reported to reduce N2O emissions by 9%e95% (Snyder et al., 2009; Bhatia et al., 2010). By analyzing data collected from 85 field studies worldwide, Akiyama et al. (2010) concluded that nitrification inhibitors significantly reduced N2O emissions by an average of 38% compared with conventional fertilizers, and their effectiveness is consistent across the range of inhibitor types, fertilizer types, and land uses investigated in this review. Nevertheless, the conclusions of this review are difficult to extrapolate globally because most of the soils included in this analysis are not typical of cropping systems in the world. In US field studies, Halvorson et al. (2010a) observed a 35% reduction in N2O by application of nitrification inhibitor combined with urea ammonium nitrate compared with application of urea ammonium nitrate alone, whereas little or no impact of nitrification inhibitors on N2O emissions was found in other studies (Parkin and Hatfield, 2010; Venterea et al., 2011). In a no-till corn-dry bean rotation system, the application of nitrification inhibitors decreased N2O emissions by 54% compared with no inhibitor (Halvorson et al., 2010b). Urease inhibitors are not as effective as polymer-coated fertilizers and nitrification inhibitors in reducing N2O emissions (Akiyama et al., 2010), probably because the hydrolysis of urea is not directly related to the production of N2O. However, the application of urease inhibitors does delay the formation of NH4 þ in soil and helps reduce N2O production derived from ammonia oxidation and subsequent denitrification of NO3 , as long as the release of NH4 þ is in synchrony with plant NH4 þ uptake. In the review of Akiyama et al. (2010), only eight published studies on the effectiveness of urease inhibitors in reducing N2O were investigated; additional data across various soil types, climates regimes, and land uses are needed to form more reliable and widely applicable conclusions. As previously indicated, the most widely used N fertilizers in agricultural production worldwide are ammonium-based (Glibert et al., 2006; USDA, 2010), thus the use of nitrification inhibitors holds potential to improve crop N recovery and mitigate soil N2O productions. By slowing nitrification, lower amounts of N2O produced through ammonia oxidation pathways and heterotrophic denitrification are anticipated due to the ability of nitrification inhibitors to diminish the effect of N concentration on N2O production and increase the crop uptake of NH4 þ (Zhu et al., 2014b). Nitrification inhibitors can be used with both synthetic and organic N fertilizers, making them a favorable option for mitigation practice. However, some studies have noted that nitrification inhibitors used with ammonium sulfate and anhydrous ammonia only reduce N2O emissions during a short period, with no effect on the overall amount of annual emissions (Parkin and Hatfield, 2010). Although consistent yields and increases in plant N uptake have been reported in crops grown with nitrification inhibitors (Hatfield and Parkin, 2014), the impact on N2O emissions is not always certain. Fertilizer source, timing, and placement, along with edaphic factors such as

Irrigation

155

temperature, pH, and water content may impact the efficacy of nitrification inhibitors in reducing N2O emissions (Kyveryga et al., 2004; Parkin and Hatfield, 2010; Hatfield and Parkin, 2014).

Consideration of Multiple Issues for Mitigation Synchronizing available soil N and crop requirements contributes to conservation of N in the soil, favors more efficient plant uptake, and therefore results in less N2O production. This can be done by optimizing in-season fertilizer timing (i.e., applying fertilizer close to when crop demand is highest), by optimizing sources (e.g., using slow-release fertilizers or changing between common sources such as anhydrous ammonia and urea), and by optimizing placement (e.g., applying fertilizer near the root zone rather than broadcasting on the soil surface). Studies have found that when N fertilizer is applied in the spring, N2O emissions are lower than the same amount of fertilizer applied in fall (Hao et al., 2001; Hultgreen, 2003), likely because crop demand for N is generally high during the growth stage but drops sharply as the crop nears maturity. Using slow-release, controlled-release, or stabilized fertilizer to enhance crop NUE and minimize N losses has also shown some promise for reducing N2O emissions, as has the use of coated and urea-based slow-release fertilizers (Snyder et al., 2009; Akiyama et al., 2010; Halvorson et al., 2010a; Hyatt et al., 2010). However, very few studies have investigated long-term effects of these fertilizers on N2O emissions. For example, the use of urea as opposed to anhydrous ammonia may tend to decrease N2O emissions, but changes in crop type and climatic conditions, as well as fertilizer rate, placement, and tillage could potentially reverse this result. Improved placement of N fertilizer, such as banding of fertilizer along crop rows, can increase the success of crop competition with microbes for available N and therefore reduce N2O production. However, the local concentration of N fertilizer in a banding application is generally high, which could potentially promote N2O production and be counterproductive to the other benefits of banding. Thus, rate modification is necessary under this type of placement with these concerns in mind. Studies on the depth of fertilizer application have shown contradictory impacts on N2O flux. Overall, the optimization of N management to mitigate climate change still relies on additional research to fully understand the complex interactions of environmental conditions affecting N2O production such as soil type and crop type with fertilizerrelated variables such as N rate, timing, source, and placement.

IRRIGATION Irrigation practices influence GHG emissions through their direct effect on soil N2O production and also include emissions from fossil fuel used to transport and deliver water. This section focuses on direct N2O emissions from soil in response to irrigation practices. Flood irrigation, furrow irrigation, (micro-) sprinkler irrigation, surface/subsurface drip irrigation, and subirrigation are the most common irrigation practices adopted in irrigated cropping systems.

156


As soil water content approaches saturation, O2 diffusion is limited, resulting in anoxic conditions that promote N2O production through both heterotrophic denitrification and ammonia oxidation pathways (see the section on Biological processes that produce N2O in soils) (Zhu et al., 2013a). However, only trace amounts of N2O can be emitted to the atmosphere once the soil is completely flooded because in fully saturated soils that are depauperate in O2, any N2O produced is likely to remain in solution and further reduced to N2 before diffusing to the atmosphere (Davidson, 1991; Dunfield et al., 1995). The optimum soil water content for N2O emissions via denitrification has been observed to exist at a water-filled pore space of 70%e90%, whereas it is assumed that nitrification produces N2O at lower water contents (Dobbie et al., 1999; Bateman and Baggs, 2005; Venterea et al., 2010). Nevertheless, Zhu et al. (2013a) observed that even under atmospheres with very little O2 (i.e., 0.5% O2), a substantial amount of N2O is still derived from ammonia oxidation pathways, indicating that ammonia oxidation can be a significant source of N2O production in soils with high water contents. Regardless of management details, most irrigation practices subject the soil to cycles of wetting and drying, which stimulate microbial activity and can result in large intermittent pulses of N2O from both denitrification and ammonia oxidation (Kieft et al., 1987; Rudaz et al., 1991; Fierer and Schimel, 2002). Flood irrigation is a practice in which an entire field is covered with water, while furrow irrigation adds water through furrows adjacent to crop beds. In flood irrigation systems, soils experience highly anoxic periods which promote denitrification processes. As discussed earlier, however, high denitrification rates do not necessarily result in high N2O emissions, as extremely anoxic conditions lead to complete denitrification (Firestone et al., 1979; Bonin et al., 1989). Field studies have supported this assumption with observations of less N2O emitted from the continuously flooded plots than from intermittently flooded plots (Katayanagi et al., 2012; Xu et al., 2013). In furrow irrigation systems, conditions approaching saturation are temporarily sustained for one or more days after irrigation. This usually leads to a large pulse of N2O and CO2 after the first irrigation event, followed by two or more pulses in the subsequent irrigation events (Davidson, 1992; Fierer and Schimel, 2002). The frequency and volume of furrow irrigations significantly impacts such cycles of soil wetting and drying and therefore N2O emissions (Davidson, 1992). The spatial variability of N2O emissions in furrow irrigation systems can also be very high. For example, furrows had significantly higher N2O emissions than beds in tomato and cotton cropping systems (Grace et al., 2010; Burger and Horwath, 2012). Micro- to landscape-scale heterogeneity in environmental conditions contributes to multiscale variability in N2O emissions (Yates et al., 2006; Heńault et al., 2012). This spatial and temporal heterogeneity in environmental conditions and flux rates makes it difficult to quantify N2O emissions from furrow irrigation systems with high levels of accuracy and precision. Compared with flood and furrow irrigations, which generate large extremes in wetting and drying soils, low-volume irrigation systems such as

Irrigation

157

surface/subsurface drip and microsprinkler irrigation usually have lower overall N2O fluxes (Nelson and Terry, 1996; Kallenbach et al., 2010; Burger and Horwath, 2012). Sprinkler systems deliver water to vegetation and the soil through a top-down approach, while surface drip irrigation supplies water from drip lines placed adjacent to crop rows. Subsurface drip irrigation targets water delivery directly to the root zone using buried pipes and tubing. These systems are typically more efficient than flood or furrow irrigation systems. For example, in a subsurface drip irrigation system, soil water contents higher than 60% of water-filled pore space have only been found within a few centimeters of the subsurface drip tape; the soil surface stays dry with a low soil water content of 20%e30% of water-filled pore space (Kallenbach et al., 2010). Consequently, evaporation losses are minimized and saturation of the whole soil profile is rare (Hanson et al., 2000; Hanson and May, 2007). In addition to reductions in evaporative water loss, fertilizer use efficiency is increased and N2O emissions are reduced using drip systems for delivering fertilizer, or fertigation. Kallenbach et al. (2010) showed a reduction in N2O emissions of 1.43 kg N/ha yr in a subsurface drip irrigation system compared to furrow irrigation. In a meta-analysis of all the available data from studies conducted in Mediterranean climates, Aguilera et al. (2013) reported that cumulative N2O emissions for furrow irrigation and drip irrigation systems were 7.8 and 2.0 kg N2O-N/ha yr, respectively. The impacts of sprinkler irrigation and surface drip irrigation on N2O fluxes are expected to be similar, although Smart et al. (2011) and Alsina et al. (2013) observed a reduction in N2O emissions during a fertigation event with a microsprinkler system compared to a surface drip irrigation system in an almond orchard. As soil moisture in subsurface irrigation systems shows less temporal variation compared with sprinkler and surface drip systems, fewer N2O pulses are expected from subsurface irrigation systems. However, side-by-side comparisons of subsurface drip irrigation and microsprinkler/surface drip irrigation are lacking. Subirrigation is an irrigation practice used in areas with relatively high water tables or where the water table can be artificially raised to allow the soil to be moistened from below the root zone. Since water is supplied to roots from below, evaporative losses are not enhanced, as they would be with surface irrigation systems. However, this system may increase N2O emissions by saturating the soil profile (Elmi et al., 2003; Munoz et al., 2005). Efficient irrigation practices, such as subsurface drip irrigation, reduce the total amount of water applied and optimize water distribution to the root zones. The reduction of soil N2O emissions through subsurface drip irrigation has been reported as a promising mitigation strategy (Burger et al., 2005; Kallenbach et al., 2010). Although it is commonly observed that irrigation events are followed by higher N2O emissions, once soil water content drops below 60% of water-filled pore space after an irrigation event, N2O emissions declines significantly (Burger et al., 2005). Kallenbach et al. (2010) also observed that the conversion from furrow irrigation to subsurface drip in cover-cropped systems decreased N2O emission by 20%e70% compared to furrow irrigation. Other

158


efficient irrigation techniques such as drip irrigation, which requires 25%e72% less water than furrow irrigation in crops with no negative yield impact (Camp, 1998; Halvorson et al., 2008), can also be used as effective N2O mitigation practices (Aguilera et al., 2013).

TILLAGE Tillage is a fundamental aspect of agricultural management that changes soil either physically, chemically, mechanically, or biologically to create suitable conditions for seedling germination and plant growth. Tillage practices are generally classified as conventional (full tillage), conservation (reduced tillage), and no-till. Conventional tillage represents the greatest level of disturbance and includes one or more passes with the following possible tillage implements: moldboard plow, disk plow, disk chisel, twisted point chisel plow, heavy-duty offset disk, subsoil chisel plow, and bedder or disk ripper. Systems with other tillage practices, such as a single pass with a ridge till implement, mulch till, or chisel plow, lead to intermediate disturbance of the soil and are classified as conservation tillage (Culman et al., 2014). No-till management is characterized by the use of seed drills and fertilizer or pesticide applicators with no additional disturbance events or implements. The influence of tillage practices on N2O emissions includes (1) fossil fuel burning during field operations; (2) their influence on soil carbon emissions and vertical distribution of carbon and total soil carbon stocks (Paustian et al., 1997); and (3) direct impacts on soil N2O emissions. This section focuses on the effect of tillage practices on N2O emissions from soil. Tillage practices influence soil N2O emissions mainly through changes in soil aeration status (Linn and Doran, 1984; Six et al., 2004; Munkholm et al., 2016) and microbial activity (Rice and Smith, 1983; Broder et al., 1984; Smith et al., 2010), which is directly related to N2O production through heterotrophic denitrification and ammonia oxidation processes (Zhu et al., 2013a). The effect of conservation and no-till relative to conventional tillage on soil N2O emissions has received much attention, but the results vary from increases to decreases or no changes (Venterea et al., 2005, 2011; Grandy et al., 2006; Mosier et al., 2006; Rochette, 2008; Kong et al., 2009; Lee et al., 2009; Halvorson et al., 2010b; Garland et al., 2011). The wide spectrum in results can be attributed to differences in the effect of climate, duration of tillage practices, N fertilizer management, and soil texture on N2O production and consumption pathways. In a 2-year tillage and crop residue study, soil denitrification rates were similar among conventional tillage, no-till, and conservation tillage (Elmi et al., 2003). However, N2O consumption rates under different soil conditions, including fluctuations in soil moisture caused by tillage, might result in varying amounts of net N2O emissions among different tillage practices. Rochette (2008) compared N2O emissions from 25 field studies managed with conventional tillage and no-till practices and concluded that the influence of tillage on N2O emissions depends on soil texture: no-till

Tillage

159

increases N2O emissions in poorly drained fine-textured soils but does not lead to any changes in well-drained soils. In a study of corn systems in Quebec, Almaraz et al. (2009) observed that changing management from conventional tillage to no-till in heavy soils increases N2O emissions. Additional studies have reported that converting from conventional tillage to no-till or conservation tillage increases N2O emissions (Almaraz et al., 2009; Kong et al., 2009; Venterea et al., 2011; Abdalla et al., 2013). However, lower N2O emissions from no-till compared to conventional tillage have been observed in humid climates after adopting no-till practices for a certain amount of time (Six et al., 2004). The overall impact of no-till management on N2O emissions is therefore highly dependent on the duration of tillage adoption (Kessavalou et al., 1998; Six et al., 2004). Conservation tillage and no-till have been widely studied for climate change mitigation, which consists of reductions in soil N2O as well as SOC sequestration (Six et al., 2004; Lee et al., 2006; Halvorson et al., 2010b; Garland et al., 2011; Venterea et al., 2011; Ogle et al., 2012; Garcia-Franco et al., 2015; Sheehy et al., 2015). The response of soil N2O emissions to reduced tillage is variable, with some systems (e.g., humid, poorly aerated, or clay soils) generally showing large increases in N2O emissions after implementation of no-till (Rochette, 2008) and other systems (e.g., dry or well-aerated soils) showing decreases or no significant response (Grandy et al., 2006). By analyzing available data for soil-derived N2O as affected by tillage under humid and dry temperate climates, Six et al. (2004) noted generally higher N2O emissions in the initial years following adoption of no-till, but after a no-till system had been in place for 10 years or more, N2O emissions were lower compared to systems under conventional tillage. Therefore, the transition from conventional to conservation tillage for the purpose of soil N2O mitigation should be considered in context of specific climates, soils, and times (Abdalla et al., 2013). The interaction between tillage practices and fertilizer types can be important in controlling N2O emissions from soils. For example, Venterea et al. (2005) reported that higher N2O emissions were found under no-till and conservation tillage compared to conventional tillage when the fertilizer was applied as urea, while the opposite was observed when anhydrous ammonia was applied, and no changes were seen with urea ammonium nitrate. Halvorson et al. (2010b) observed that no difference was found between conventional tillage and no-till when no fertilizer was applied, whereas conventional tillage significantly increased N2O emissions compared to no-till when urea was applied. Conventional tillage has also been shown to diminish the effect of enhancedefficiency N fertilizer on N2O emissions compared to no-till practices, in which the use of enhanced-efficiency N fertilizer led to a significant decline in N2O emissions (Halvorson et al., 2010b). In addition, fertilizer placement is another important factor influencing N2O emissions under different tillage practices (Kessel et al., 2013). Venterea and Stanenas (2008) and Venterea et al. (2005) indicated that deep N fertilizer placement was likely to increase N2O emissions when accompanied by conventional tillage as compared to no-till.

160


COVER CROPS AND ORGANIC AMENDMENTS The release of nutrients from organic amendments into forms that can be taken up by plants is generally slower compared to synthetic fertilizer. Thus, less leaching and nutrient loss and ultimately greater nutrient use efficiencies are expected for organic fertilizers compared to mineral fertilizers (Crews and Peoples, 2004; Seiter and Horwath, 2004; Drinkwater and Snapp, 2007). Cover crops and organic amendments are the most common ways to increase soil organic fertility: cover crops as a special type of organic amendments include numerous crops that are typically planted, grown, and not harvested in the field but instead returned to the soil through mowing or tillage; organic amendments include any types of plant- or animal-based residue, e.g., manure, compost, or biosolids (Culman et al., 2014). The planting time and functionality of cover crops vary according to cropping system and climate (Snapp et al., 2005). For example, winter legumes are planted in winter to fix N and build soil organic matter (Kallenbach et al., 2010), rye is planted in the fall to suppress weeds, and sorghum sudangrass is planted in the summer to break up compacted soil (Shipley et al., 1992; Wolfe, 1997; Matthiessen and Kirkegaard, 2006). The use of cover crops generally tends to increase N2O production in soil (Steenwerth and Belina, 2008; De Gryze et al., 2009; Kallenbach et al., 2010; Smukler et al., 2012). Cover crops, especially N-rich legumes, can add a substantial amount of C and N to the soil, thereby decreasing aeration due to microbial respiration and increasing microbial activity responsible for N2O production (Varco et al., 1987; Aulakh et al., 1991; Follett, 2001; Watson et al., 2002; Christopher and Lal, 2007; Sainju et al., 2007). Soil N2O production can also increase if the N released from cover crops is not well synchronized with the needs of subsequent crops. Here, excess N and ample C in soil can increase N2O derived from heterotrophic denitrification and ammonia oxidation pathways (Smid and Beauchamp, 1976; Firestone et al., 1979; Stark and Firestone, 1995; Zhu et al., 2013a). However, the effect of cover crops on soil N2O production varies depending on the type of cover crop used. For example, application of leguminous residue may introduce and increase available soil N and therefore lead to increases in N2O emissions, whereas a winter grass may rapidly take up surplus N in soil after harvest and have a greater chance of reducing N2O emissions (Culman et al., 2014). Future studies focused on empirical data collection are needed to support this hypothesis. The common use of organic amendments such as manure and compost in soil management has been increasingly suggested as a means to recouple C and N cycling in agricultural ecosystems to improve their sustainability (Gardner and Drinkwater, 2009). Application of organic amendments can promote increased water holding capacity, C availability, and tilth and provide important crop nutrients (Celik et al., 2004; Leroy et al., 2008; Wright et al., 2008; Zhu et al., 2013b). However, the effect of a given organic amendment on N2O emissions after its application to soils varies widely, from decreasing

Cover Crops and Organic Amendments

161

N2O emissions (Meijide et al., 2007; Alluvione et al., 2010; Dalal et al., 2010; Vaughan et al., 2011) to increasing emissions (Zaman et al., 2004; Rodriguez et al., 2011; Zhu et al., 2013b). The application of organic amendments may influence N2O emissions by impacting the availability of C and N, which subsequently may determine the predominant N2O production pathway; the main or combined mechanisms and pathways of N2O production, however, depends on the characteristics of the amendment, soil conditions, and climate. For example, application of green waste compost in sandy soils increased N2O emissions compared to the application of mineral N fertilizer, while no differences were found in clay soils (Zhu-Barker et al., 2015b). Vaughan et al. (2011) reported that the application of green waste compost decreased N2O emissions compared to the application of mineral N fertilizer, whereas the application of fresh green waste did not change emissions. Dalal et al. (2010) observed higher N2O emissions in soil amended with feedlot manure compared to soil amended with urea, while application of green waste compost lowered emissions. Carbon inputs likely affect O2 availability, which directly affects N2O production from both ammonia oxidation and heterotrophic denitrification (Zhu et al., 2013a). Therefore, application of organic amendments can affect soil N2O productions by changing the availability and spatial distribution of O2 in soil (Zhu et al., 2014a). Other than synthetic N fertilizers, the application of organic amendments, especially manure, in agricultural lands is a main contributor to increased N2O in atmosphere (Davidson, 2009). The cobenefits for soil C sequestration, improvements in soil fertility, and slow mineralization of organic N combined with N2O mitigation potential has led to growing interest in the use of organic amendments. Recent work by Bowles et al. (2015) has shown in organic tomato production systems in California, fields that received only organic matter inputs and supported high labile C and N content had low soil inorganic N pools and, thus, tightly coupled plant-soil N cycling with low potential for N loss. Tomato yields were maintained at approximately 80% of the county average, suggesting that continued work is needed to achieve yields of fields receiving synthetic N fertilizer. Considering that N2O emissions from managed manure in the United States were estimated to range from 2.6 to 30.6 Mt CO2e/yr (USEPA, 2009) and that most manure was applied in corn systems in the United States, improved manure management in corn cropland alone could lead to significant mitigation (MacDonald, 2009). Generally, emissions of N2O after manure application depend on manure type, climatic conditions, and water content in both the soil and manure. A number of studies have proposed manure management practices to reduce N2O emissions. For example, Paustian et al. (2004) and Saggar et al. (2004) noted that manure application to coincide with dry soil and low temperature could reduce N2O emissions, while Greogorich et al. (2005) proposed that the application of solid manure rather than liquid manure is preferable to reduce emissions. Reduction in the application rate of pig slurry has also

162


been reported to reduce N2O emissions (Rochette et al., 2000). Nonetheless, the most promising practice to reduce soil N2O emissions through nutrient management may be adjusting the synthetic N application rates to account for N availability in the manure. This adjustment, however, is not easy because it requires a full understanding of the N mineralization characteristics of manure during the period of crop growth. So far, less than 60% of farmers in the United States have made such adjustments (MacDonald, 2009).

CLIMATE CHANGE AND SOIL N2O PRODUCTION Climate change, manifest as increasing temporal and spatial variability in temperature, precipitation, and winds, particularly the incidence and magnitude of extreme events, can cause significant changes in agricultural systems (Fig. 6.6) (Reilly et al., 1996; IPCC, 2007a; Morton, 2007; Reidsma et al., 2010; Vermeulen et al., 2012). For example, projected increases in temperatures and heat waves and in water shortages have been cited to potentially reduce crop yields and area suitable for cropping (Ciais et al., 2005; Lee et al., 2008; Reidsma et al., 2010; Lobell et al., 2011), which will in turn affect the livelihood of farmers as fewer opportunities are available for agricultural expansion on additional lands (Schro¨ter et al., 2005; Metzger et al., 2006). Rising atmospheric CO2 concentrations may increase net primary productivity and soil C content while at the same time decreasing food and forage quality (Parton et al., 1995; Paustian et al., 1995; Taub et al., 2008; Myers et al., 2014). Soil N pools can also diminish in response to changes in the frequency and magnitude of dry/ hot scenarios (Dueri et al., 2007). Agriculture is facing the challenge of producing more on the same amount of land while facing pressures to adapt to a changing climate and become more resilient to climate variability. This challenge is great, given the continuous and increasing demand for agricultural products due to population growth, changes in diet related to increased per capita income,

FIGURE 6.6 Proposed conceptual feedback effect of climate change on soil N2O production.

Climate Change and Soil N2O Production

163

finite quantities of suitable land available for agricultural expansion, and competing entities like biofuel production for this same land area (Steenwerth et al., 2014). As a result, land management practices need to be adjusted to maintain agricultural crop production under changing climate; reducing soil N2O emissions is an important component of mitigating climate change through improved land management (Montzka et al., 2011; Xu-Ri et al., 2012). Changes in climate, including increased temperature and frequency of heat waves and heavy precipitation, dramatically influence soil conditions such as water content, temperature, and microbial activity, and therefore ultimately influence N2O production (Fig. 6.6) (Lu and Cheng, 2009; Abdalla et al., 2010; Singh et al., 2010; Reay et al., 2012; Kessel et al., 2013). Dueri et al. (2007) reported that the amount of soil N loss increased when a farm model was run under scenarios of more warmewet or hotedry conditions. Lu and Cheng (2009) also observed that N2O emissions from soil significantly increased as temperature and precipitation increased. As discussed in the section on Biological production of N2O in soils, soil N2O production can increase significantly when heterotrophic denitrification and ammonia oxidation pathways are stimulated by increased soil temperature and moisture (Keeney et al., 1979; Avrahami et al., 2003; Szukics et al., 2010). Climate change, e.g., increases in temperatures and heat waves, tends to reduce the long-term crop yields by damaging crops at particular developmental stages (Lobell and Field, 2007; Lee et al., 2008; Lobell et al., 2011). For example, global data indicate that wheat, maize, and barley have experienced climate warmingeassociated yield reductions of 40 Mt/yr since1981 (Lobell and Field, 2007). Lobell et al. (2011) analyzed data from historical maize trials in Africa and also showed that increased heat (daily temperatures >30 C) reduced yield by 1% on average in rain-fed maize systems. Suboptimal crop growth under such conditions could leave large amounts of residual fertilizer N in the soil to be used by soil microorganisms to produce N2O (Burger and Horwath, 2012; Culman et al., 2014). As a result, the emission intensity of a system (N2O emissions per unit yield) may increase, contributing to future global climate change. Changes in climate will also determine changes in land management as farmers strive to maximize profitability. Therefore, changes in soil N2O production may be a complex function of climate-driven land management changes, as opposed to a simple direct effect of either climate or land management. In many systems, management practices in agriculture will need to be developed to build resilience to evolving climatic change factors to take advantage of potential cobenefits for mitigation as a secondary goal. Therefore, future climate change will necessitate evaluations of the role of management development in determining potential responses of agroecosystems to global change, especially when considering management practices that can be adopted to mitigate soil N2O production (Montzka et al., 2011; Xu-Ri et al., 2012). Upland soils that are managed to support high crop productivity produce large quantities of N2O and contribute more N2O to the atmosphere than any other anthropogenic source (Smith et al., 2008, 2014). Reducing this source

164


represents a substantial climate change mitigation opportunity, particularly since N2O is a major source of climate change forcing in intensively managed cropland. In such systems, soil N2O emissions are directly related to N inputs. Therefore, strategies to develop better N management are the key focal point to reduce soil N2O emissions and mitigate future climate change. Improved N management practices to reduce soil N2O emissions would also limit the negative impacts of N fertilizer application such as NO3 pollution of water and NH3 volatilization (Gebbers and Adamchuk, 2010). Other practices such as conservational tillage, optimized irrigation, and improved management of organic amendments can also contribute to the reduction of soil N2O.

CONCLUSIONS This chapter summarizes known pathways that contribute to soil N2O emissions, describes the factors that control these pathways, highlights how land management practices influence soil N2O emissions by inducing changes in these factors, and emphasizes the response of soil N2O emission to climate change and potential mitigation practices. With a focus on soil N2O production pathways and their controls, this chapter has outlined the role of commonly used practices on soil N2O emissions, such as fertilization, irrigation, tillage, cover cropping, and organic amendments. Future climate change is expected to affect soil N2O emissions, and the development of land management practices to reduce emissions is necessary to abate further climate change. Important insights with respect to previously unanticipated pathways and mechanisms of soil N2O production have expanded the horizons of researchers and stimulated recent efforts to develop new methods for quantifying these pathways, but critical challenges remain. For example, N2O production processes such as nitrifier denitrification and DNRA remain difficult to quantify, and the mechanisms behind these pathways are still unknown. The variability and heterogeneity of N2O production pathways and their complex relationship with soil conditions also limits the extent to which well-constrained estimates can be made relevant to climate change issues. Future efforts to meet these challenges would constitute a comprehensive research comprised of the following elements: 1. dedicated research to continue to develop and improve techniques for the characterization, quantification, and modeling of alternative N2O production pathways; 2. state-of-the-art molecular techniques to determine the functional microbial communities involved in each pathway; and 3. consideration of changes in the ecological drivers of soil N2O productions as a result of climate change, and establishing links between microbial and ecosystem scales. Nitrogen fertilizers are a major source of soil N2O emissions in croplands as they are a direct source of substrate for different soil N2O production pathways. Reducing N application rates has the potential to reduce soil N2O emissions, but

References 165 there is a risk of lower crop yields, to the point at which overall yield-scaled emissions would be in fact higher. Splitting an N application into small frequent applications, to match in-field variations in crop N demand and soil N supply, as well as to avoid wet periods, will help avoid pulses of N2O emission. However, technologies that can meet these goals are in need of further development and cost-benefit analyses. Alternative N fertilizers, e.g., slow-release and coated fertilizers or products containing nitrification inhibitors, have shown promise in reducing soil N2O emissions. Notwithstanding, the efficiency of these products depends primarily on environmental factors and soil characteristics, and therefore synchronization of N availability with plant demand may not necessarily improve. Other land management practices centered around tillage, irrigation, cover cropping, and the application of organic amendments also influence N2O production from different pathways by changing relevant soil conditions such as water content, aeration, organic C availability, and the activity of microorganisms. Nevertheless, many interactions are possible between N fertilizer application rate, changes in timing, placement, and source, tillage intensity, irrigation efficiency, and specific properties of organic amendments. The outcome of these interactions still leaves large uncertainties when alternative management practices are considered for mitigation practices. Responding effectively to climate change requires recognition of the role of land management practices as a whole in determining how soil N2O emissions will be affected by global change. Management strategies manipulate many different variables related to N2O production, but the net outcome of a given management strategy on soil N2O emission is often not simply the cumulative effect of each component of the management strategy. Mitigation practices are being increasingly examined and employed in agriculture and will continue to improve as knowledge gaps are filled regarding the mechanistic understanding of N2O production process in soil as well as generation of quality data for evaluating and refining predictive models. These scientific advances will increase in value as they are accompanied by improved estimates of the cost efficiency of management development, greater engagement of land users (education and outreach), and greater incentive on the part of policy makers to design farseeing strategies that meet both agricultural and environmental goals.

ACKNOWLEDGMENTS We also thank Timothy A. Doane, G. Philip Robertson, associate editor R. Kelman Wieder, and four anonymous reviewers for thoughtful comments on earlier versions of this manuscript.

REFERENCES Abdalla, M., Jones, M., Williams, M., 2010. Simulation of N2O fluxes from Irish arable soils: effect of climate change and management. Biology and Fertility of Soils 46, 247e260. Abdalla, M., Osborne, B., Lanigan, G., Forristal, D., Williams, M., Smith, P., Jones, M., 2013. Conservation tillage systems: a review of its consequences for greenhouse gas emissions. Soil Use and Management 29, 199e209.

166


Aguilera, E., Lassaletta, L., Sanz-Cobena, A., Garnier, J., Vallejo, A., 2013. The potential of organic fertilizers and water management to reduce N2O emissions in Mediterranean climate cropping systems. A review. Agriculture, Ecosystems and Environment 164, 32e52. Akiyama, H., Yan, X.Y., Yagi, K., 2010. Evaluation of effectiveness of enhanced-efficiency fertilizers as mitigation options for N2O and NO emissions from agricultural soils: meta-analysis. Global Change Biology 16, 1837e1846. Alcańtara, V., Don, A., Well, R., Nieder, R., 2016. Deep ploughing increases agricultural soil organic matter stocks. Global Change Biology 22, 2939e2956. Alluvione, F., Bertora, C., Zavattaro, L., Grignani, C., 2010. Nitrous oxide and carbon dioxide emissions following green manure and compost fertilization in corn. Soil Science Society of America Journal 74, 384e395. Almaraz, J.J., Mabood, F., Zhou, X.M., Madramootoo, C., Rochette, P., Ma, B.L., Smith, D.L., 2009. Carbon dioxide and nitrous oxide fluxes in corn grown under two tillage systems in Southwestern Quebec. Soil Science Society of America Journal 73, 113e119. Alsina, M.M., Fanton-Borges, A.C., Smart, D.R., 2013. Spatiotemporal variation of event related N2O and CH4 emissions during fertigation in a California almond orchard. Ecosphere 4, 1e21. Anderson, I.C., Levine, J.S., 1986. Relative rates of nitric-oxide and nitrous-oxide production by nitrifiers, denitrifiers, and nitrate respirers. Applied and Environmental Microbiology 51, 938e945. Anderson, I.C., Poth, M., Homstead, J., Burdige, D., 1993. A comparison of NO and N2O production by the autotrophic nitrifier nitrosomonas-europaea and the heterotrophic nitrifier alcaligenes-faecalis. Applied and Environmental Microbiology 59, 3525e3533. Andersson, K.K., Hooper, A.B., 1983. O2 and H2O are each the source of one O in NO2 produced from NH3 by Nitrosomonas - N-15-Nmr evidence. FEBS Letters 164, 236e240. Anthonisen, A., Loehr, R., Prakasam, T., Srinath, E., 1976. Inhibition of nitrification by ammonia and nitrous acid. Journal (Water Pollution Control Federation) 835e852. Arah, J., 1997. Apportioning nitrous oxide fluxes between nitrification and denitrification using gas-phase mass spectrometry. Soil Biology and Biochemistry 29, 1295e1299. Aulakh, M., Doran, J., Walters, D., Power, J., 1991. Legume residue and soil water effects on denitrification in soils of different textures. Soil Biology and Biochemistry 23, 1161e1167. Avrahami, S., Bohannan, B.J.A., 2007. Response of Nitrosospira sp strain AF-Like ammonia oxidizers to changes in temperature, soil moisture content, and fertilizer concentration. Applied and Environmental Microbiology 73, 1166e1173. Avrahami, S., Conrad, R., Braker, G., 2002. Effect of soil ammonium concentration on N2O release and on the community structure of ammonia oxidizers and denitrifiers. Applied and Environmental Microbiology 68, 5685e5692. Avrahami, S., Liesack, W., Conrad, R., 2003. Effects of temperature and fertilizer on activity and community structure of soil ammonia oxidizers. Environmental Microbiology 5, 691e705. Avrhami, S., Conrad, R., 2003. Patterns of community change among ammonia oxidizers in meadow soils upon long-term incubation at different temperatures. Applied and Environmental Microbiology 69, 6152e6164. Azam, F., Muller, C., Weiske, A., Benckiser, G., Ottow, J.C.G., 2002. Nitrification and denitrification as sources of atmospheric nitrous oxide - role of oxidizable carbon and applied nitrogen. Biology and Fertility of Soils 35, 54e61. Barraclough, D., 1995. The use of 15N pool dilution and enrichment to separate the heterotrophic and autotrophic pathways of nitrification. Soil Biology and Biochemistry 27, 6. Bateman, E.J., Baggs, E.M., 2005. Contributions of nitrification and denitrification to N2O emissions from soils at different water-filled pore space. Biology and Fertility of Soils 41, 379e388. Beaulieu, J.J., Tank, J.L., Hamilton, S.K., Wollheim, W.M., Hall, R.O., Mulholland, P.J., Peterson, B.J., Ashkenas, L.R., Cooper, L.W., Dahm, C.N., 2011. Nitrous oxide emission from denitrification in stream and river networks. Proceedings of the National Academy of Sciences of the United States of America 108, 214e219.

References 167 Berg, P., Klemedtsson, L., Rosswall, T., 1982. Inhibitory effect of low partial pressures of acetylene on nitrification. Soil Biology and Biochemistry 14, 301e303. Betlach, M.R., Tiedje, J.M., 1981. Kinetic explanation for accumulation of nitrite, nitric oxide, and nitrous oxide during bacterial denitrification. Applied and Environmental Microbiology 42, 1074e1084. Bhatia, A., Sasmal, S., Jain, N., Pathak, H., Kumar, R., Singh, A., 2010. Mitigating nitrous oxide emission from soil under conventional and no-tillage in wheat using nitrification inhibitors. Agriculture, Ecosystems and Environment 136, 247e253. Blagodatsky, S., Kesik, M., Papen, H., Butterbach-Bahl, K., 2006. Production of NO and N2O by the heterotrophic nitrifier Alcaligenes faecalis parafaecalis under varying conditions of oxygen saturation. Geomicrobiology Journal 23, 165e176. Bock, E., Koops, H.-P., Harms, H., 1986. Cell Biology of Nitrifying Bacteria. Bollag, J.-M., Tung, G., 1972. Nitrous oxide release by soil fungi. Soil Biology and Biochemistry 4, 271e276. Bollmann, A., Conrad, R., 1998. Influence of O2 availability on NO and N2O release by nitrification and denitrification in soils. Global Change Biology 4, 387e396. Bonin, P., Gilewicz, M., Bertrand, J.C., 1989. Effects of oxygen on each step of denitrification on Pseudomonas-Nautica. Canadian Journal of Microbiology 35, 1061e1064. Bouman, O., Curtin, D., Campbell, C., Biederbeck, V., Ukrainetz, H., 1995. Soil acidification from long-term use of anhydrous ammonia and urea. Soil Science Society of America Journal 59, 1488e1494. Bouwman, A., Boumans, L., Batjes, N., 2002a. Emissions of N2O and NO from fertilized fields: summary of available measurement data. Global Biogeochemical Cycles 16. Bouwman, A.F., Boumans, L.J.M., Batjes, N.H., 2002b. Modeling global annual N2O and NO emissions from fertilized fields. Global Biogeochemical Cycles 16, 1e8. Bowles, T.M., Hollander, A.D., Steenwerth, K., Jackson, L.E., 2015. Tightly-coupled plantsoil nitrogen cycling: comparison of organic farms across an agricultural landscape. PLoS One 10, e0131888. Braker, G., Conrad, R., 2011. Diversity, structure, and size of N2O-Producing microbial communities in soilsdwhat matters for their functioning? Advances in Applied Microbiology 75. Bray, W.C., Simpson, M.E., MacKenzie, A.A., 1919. The volumetric determination of hydroxylamine. Journal of the American Chemical Society 41, 1363e1378. Breitenbeck, G.A., Blackmer, A.M., Bremner, J.M., 1980. Effects of different nitrogen fertilizers on emission of nitrous-oxide from soil. Geophysical Research Letters 7, 85e88. Breitenbeck, G.A., Bremner, J.M., 1986. Effects of rate and depth of fertilizer application on emission of nitrous-oxide from soil fertilized with anhydrous ammonia. Biology and Fertility of Soils 2, 201e204. Bremner, J., Blackmer, A., Waring, S., 1980. Formation of nitrous oxide and dinitrogen by chemical decomposition of hydroxylamine in soils. Soil Biology and Biochemistry 12, 263e269. Bremner, J., Shaw, K., 1958. Denitrification in soil. I. Methods of investigation. The Journal of Agricultural Science 51, 22e39. Bremner, J.M., 1997. Sources of nitrous oxide in soils. Nutrient Cycling in Agroecosystems 49, 7e16. Broder, M., Doran, J., Peterson, G., Fenster, C., 1984. Fallow tillage influence on spring populations of soil nitrifiers, denitrifiers, and available nitrogen. Soil Science Society of America Journal 48, 1060e1067. Buresh, R.J., Moraghan, J.T., 1976. Chemical reduction of nitrate by ferrous iron. Journal of Environmental Quality 5, 320e325. Burford, J., Bremner, J., 1975. Relationships between the denitrification capacities of soils and total, water-soluble and readily decomposable soil organic matter. Soil Biology and Biochemistry 7, 389e394. Burger, M., Horwath, W., 2012. Assessment of Baseline Nitrous Oxide Emissions in California Cropping Systems. California Air Resources Board.

168


Burger, M., Jackson, L.E., Lundquist, E.J., Louie, D.T., Miller, R.L., Rolston, D.E., Scow, K.M., 2005. Microbial responses and nitrous oxide emissions during wetting and drying of organically and conventionally managed soil under tomatoes. Biology and Fertility of Soils 42, 109e118. Burger, M., Venterea, R.T., 2011. Effects of Nitrogen Fertilizer Types on Nitrous Oxide Emissions. In: Understanding Greenhouse Gas Emissions from Agricultural Management, 1072, pp. 179e202. Burth, I., Ottow, J., 1983. Influence of pH on the production of N₂O and N₂ by different denitrifying bacteria and Fusarium solani. Ecological Bulletins 207e215. Burton, D., Li, X., Grant, C., 2008a. Influence of fertilizer nitrogen source and management practice on N2O emissions from two Black Chernozemic soils. Canadian Journal of Soil Science 88, 219e227. Burton, D., Zebarth, B., Gillam, K., MacLeod, J., 2008b. Effect of split application of fertilizer nitrogen on Nw 2O emissions from potatoes. Canadian Journal of Soil Science 88, 229. Butt, W., Lees, H., 1960. The biochemistry of the nitrifying organisms. 6. The effect of oxygen concentration on nitrite oxidation in the presence of different inorganic ions. Biochemical Journal 76, 425. Camp, C., 1998. Subsurface drip irrigation: a review. Transactions of the ASAE 41, 1353. Caskey, W.H., Tiedje, J.M., 1979. Evidence for clostridia as agents of dissimilatory reduction of nitrate to ammonium in soils. Soil Science Society of America Journal 43, 931e936. CDFA, 2012. Fertilizer Tonnage Reports. California Department of Food and Agriculture, Sacramento, CA. Celik, I., Ortas, I., Kilic, S., 2004. Effects of compost, mycorrhiza, manure and fertilizer on some physical properties of a Chromoxerert soil. Soil and Tillage Research 78, 59e67. Chalk, P., Keeney, D., Walsh, L., 1975. Crop recovery and nitrification of fall and spring applied anhydrous ammonia. Agronomy Journal 67, 33e37. Chalk, P.M., Smith, C.J., 1983. Chemodenitrification. In: Freney, J.R., Simpson, J.R. (Eds.), Gaseous Loss of Nitrogen from Plant-Soil Systems. Development in Plant and Soils Sciences, vol. 9, p. 25. Chao, T.-T., Kroontje, W., 1966. Inorganic nitrogen transformations through the oxidation and reduction of iron. Soil Science Society of America Journal 30, 193e196. Chao, Y., Mao, Y., Yu, K., Zhang, T., 2016. Novel nitrifiers and comammox in a full-scale hybrid biofilm and activated sludge reactor revealed by metagenomic approach. Applied Microbiology and Biotechnology 1e13. Chapuis-Lardy, L., Wrage, N., Metay, A., Chotte, J.L., Bernoux, M., 2007. Soils, a sink for N(2)O? A review. Global Change Biology 13, 1e17. Christopher, S.F., Lal, R., 2007. Nitrogen management affects carbon sequestration in North American cropland soils. Critical Reviews in Plant Sciences 26, 45e64. Ciais, P., Reichstein, M., Viovy, N., Granier, A., Ogeé, J., Allard, V., Aubinet, M., Buchmann, N., Bernhofer, C., Carrara, A., 2005. Europe-wide reduction in primary productivity caused by the heat and drought in 2003. Nature 437, 529e533. Cle´ment, J.C., Shrestha, J., Ehrenfeld, J.G., Jaffe, P.R., 2005. Ammonium oxidation coupled to dissimilatory reduction of iron under anaerobic conditions in wetland soils. Soil Biology and Biochemistry 37, 2323e2328. Cole, J., 1988. Assimilatory and dissimilatory reduction of nitrate to ammonia. Symposium of the Society for General Microbiology 281e329. Cole, J., Brown, C., 1980. Nitrite reduction to ammonia by fermentative bacteria: a short circuit in the biological nitrogen cycle. FEMS Microbiology Letters 7, 65e72. Conrad, R., 1996. Soil microorganisms as controllers of atmospheric trace gases (H2, CO, CH4, OCS, N2O, and NO). Microbiological Reviews 60, 609e640. Cooper, G.S., Smith, R., 1963. Sequence of products formed during denitrification in some diverse western soils. Soil Science Society of America Journal 27, 659e662. Crews, T., Peoples, M., 2004. Legume versus fertilizer sources of nitrogen: ecological tradeoffs and human needs. Agriculture. Ecosystems and Environment 102, 279e297.

References 169 Culman, S.W., Haden, V.R., Maxwell, T., Waterhouse, H., Horwath, W., 2014. Greenhouse Gas Mitigation Opportunities in California Agriculture: Review of California Cropland Emissions and Mitigation Potential. Daims, H., Lebedeva, E.V., Pjevac, P., Han, P., Herbold, C., Albertsen, M., Jehmlich, N., Palatinszky, M., Vierheilig, J., Bulaev, A., 2015. Complete nitrification by Nitrospira bacteria. Nature 528, 504e509. Dalal, R.C., Gibson, I., Allen, D.E., Menzies, N.W., 2010. Green waste compost reduces nitrous oxide emissions from feedlot manure applied to soil. Agriculture, Ecosystems and Environment 136, 273e281. Dalal, R.C., Wang, W.J., Robertson, G.P., Parton, W.J., 2003. Nitrous oxide emission from Australian agricultural lands and mitigation options: a review. Australian Journal of Soil Research 41, 165e195. Davidson, E.A., 1991. Fluxes of nitrous oxide and nitric oxide from terrestrial ecosystems. In: Rogers, J.E., Whitman, W.B. (Eds.), Microbial Production and Consumption of Greenhouse Gases: Methane, Nitrogen Oxides and Halomethanes. American Society of Microbiology, Washington, DC, pp. 219e235. Davidson, E.A., 1992. Sources of nitric oxide and nitrous oxide following wetting of dry soil. Soil Science Society of America Journal 56, 95e102. Davidson, E.A., 2009. The contribution of manure and fertilizer nitrogen to atmospheric nitrous oxide since 1860. Nature Geoscience 2, 659e662. De Bie, M.J., Middelburg, J.J., Starink, M., Laanbroek, H.J., 2002. Factors Controlling Nitrous Oxide at the Microbial Community and Estuarine Scale. In: Marine Ecology Progress Series, vol. 240. De Boer, W., Kowalchuk, G., 2001. Nitrification in acid soils: micro-organisms and mechanisms. Soil Biology and Biochemistry 33, 853e866. De Gryze, S., Albarracin, M., Catala-Luque, R., Howitt, R., Six, J., 2009. Modeling shows that alternative soil management can decrease greenhouse gases. California Agriculture 63, 84e90. de Klein, C., Novoa, R.S., Ogle, S., Smith, K., Rochette, P., Wirth, T., McConkey, B., Mosier, A., Rypdal, K., Walsh, M., 2006. N2O emissions from managed soils, and CO2 emissions from lime and urea application. In: IPCC Guidelines for National Greenhouse Gas Inventories, Prepared by the National Greenhouse Gas Inventories Programme, vol. 4, pp. 1e54. De Klein, C., Van Logtestijn, R., 1994. Denitrification in the top soil of managed grasslands in The Netherlands in relation to soil type and fertilizer level. Plant and Soil 163, 33e44. Dickinson, R.E., Cicerone, R.J., 1986. Future global warming from atmospheric trace gases. Nature 319, 109e115. Dobbie, K.E., McTaggart, I.P., Smith, K.A., 1999. Nitrous oxide emissions from intensive agricultural systems: variations between crops and seasons, key driving variables, and mean emission factors. Journal of Geophysical Research: Atmospheres 104, 26891e26899. dos Santos, V.A.P.M., Bruijnse, M., Tramper, J., Wijffels, R.H., 1996. The magic bead concept: an integrated approach to nitrogen removal with co-immobilized microorganisms. Applied Microbiology and Biotechnology 45, 447e453. Drinkwater, L.E., Snapp, S., 2007. Nutrients in agroecosystems: rethinking the management paradigm. Advances in Agronomy 92, 163e186. Drury, C., Reynolds, W., Tan, C., Welacky, T., Calder, W., McLaughlin, N., 2006. Emissions of nitrous oxide and carbon dioxide. Soil Science Society of America Journal 70, 570e581. Dueri, S., Calanca, P., Fuhrer, J., 2007. Climate change affects farm nitrogen losseA Swiss case study with a dynamic farm model. Agricultural Systems 93, 191e214. Dunfield, P.F., Topp, E., Archambault, C., Knowles, R., 1995. Effect of nitrogen fertilizers and moisture content on CH4 and N2O fluxes in a humisol: measurements in the field and intact soil cores. Biogeochemistry 29, 199e222. Eagle, J., Henry, L.R., Olander, L.P., Haugen-Kozyra, K., Millar, N., Robertson, G.P., 2011. Greenhouse Gas Mitigation Potential of Agricultural Land Management in the United States: A Synthesis of the Literature. Technical Working Group on Agricultural Greenhouse Gases (T-AGG) Report.

170


Elmi, A.A., Madramootoo, C., Hamel, C., Liu, A., 2003. Denitrification and nitrous oxide to nitrous oxide plus dinitrogen ratios in the soil profile under three tillage systems. Biology and Fertility of Soils 38, 340e348. Engel, R., Liang, D.L., Wallander, R., Bembenek, A., 2010. Influence of urea fertilizer placement on nitrous oxide production from a silt loam soil. Journal of Environmental Quality 39, 115e125. ERS, 2016. U.S. Fertilizer Use and Price USDA Economic Research Service. http://www.ers. usda.gov/Data/FertilizerUse/. Fierer, N., Schimel, J.P., 2002. Effects of dryingerewetting frequency on soil carbon and nitrogen transformations. Soil Biology and Biochemistry 34, 777e787. Firestone, M., 1982. Biological denitrification. In: Stevenson, F.J., et al. (Eds.), Nitrogen in Agricultural Soils. Soil Science Society of America, Madison, WI, USA, p. 38. Firestone, M.K., Davidson, E.A., 1989a. Microbial basis of NO and N2O production and consumption in soil. In: Andreae, M.O., Schimel, D.S. (Eds.), Exchange of Trace Gases between Terrestrial Ecosystems and the Atmosphere. Life Sciences Research Reports. Wiley, Chichester, UK, pp. 7e21. Firestone, M.K., Davidson, E.A., 1989b. Microbiological basis of NO and N2O production and consumption. In: Andreae, M.O., Schimel, D.S. (Eds.), Exchange of Trace Gases between Terrestrial Ecosystems and the Atmosphere. Wiley, New York, pp. 7e21. Firestone, M.K., Firestone, R.B., Tiedje, J.M., 1980. Nitrous oxide from soil denitrification: factors controlling its biological production. American Association for the Advancement of Science Science (Washington DC) 208, 749e751. Firestone, M.K., Smith, M.S., Firestone, R.B., Tiedje, J.M., 1979. Influence of nitrate, nitrite, and oxygen on the composition of the gaseous products of denitrification in soil. Soil Science Society of America Journal 43, 1140e1144. Follett, R., 2001. Soil management concepts and carbon sequestration in cropland soils. Soil and Tillage Research 61, 77e92. Fortuna, A.M., Honeycutt, C.W., Vandemark, G., Griffin, T.S., Larkin, R.P., He, Z.Q., Wienhold, B.J., Sistani, K.R., Albrecht, S.L., Woodbury, B.L., Torbert, H.A., Powell, J.M., Hubbard, R.K., Eigenberg, R.A., Wright, R.J., Alldredge, J.R., Harsh, J.B., 2012. Links among nitrification, nitrifier communities, and edaphic properties in contrasting soils receiving dairy slurry. Journal of Environmental Quality 41, 262e272. Frame, C.H., Casciotti, K.L., 2010. Biogeochemical controls and isotopic signatures of nitrous oxide production by a marine ammonia-oxidizing bacterium. Biogeosciences 7, 2695e2709. Fujinuma, R., Venterea, R.T., Rosen, C., 2011. Broadcast urea reduces N2O but increases no emissions compared with conventional and shallow-applied anhydrous ammonia in a coarse-textured soil. Journal of Environmental Quality 40, 1806e1815. Gagnon, B., Ziadi, N., 2010. Grain corn and soil nitrogen responses to sidedress nitrogen sources and applications. Agronomy Journal 102, 1014e1022. Gagnon, B., Ziadi, N., Rochette, P., Chantigny, M.H., Angers, D.A., 2011. Fertilizer source influenced nitrous oxide emissions from a clay soil under corn. Soil Science Society of America Journal 75, 595e604. Galloway, J.N., Townsend, A.R., Erisman, J.W., Bekunda, M., Cai, Z.C., Freney, J.R., Martinelli, L.A., Seitzinger, S.P., Sutton, M.A., 2008. Transformation of the nitrogen cycle: recent trends, questions, and potential solutions. Science 320, 889e892. Garcia-Franco, N., Albaladejo, J., Almagro, M., Martıńez-Mena, M., 2015. Beneficial effects of reduced tillage and green manure on soil aggregation and stabilization of organic carbon in a Mediterranean agroecosystem. Soil and Tillage Research 153, 66e75. Gardner, J.B., Drinkwater, L.E., 2009. The fate of nitrogen in grain cropping systems: a metaanalysis of 15N field experiments. Ecological Applications 19, 2167e2184. Garland, G.M., Suddick, E., Burger, M., Horwath, W., Six, J., 2011. Direct N 2 O emissions following transition from conventional till to no-till in a cover cropped Mediterranean vineyard (Vitis vinifera). Agriculture, Ecosystems and Environment 144, 423e428.

References 171 Gebbers, R., Adamchuk, V.I., 2010. Precision agriculture and food security. Science 327, 828e831. Gillam, K., Zebarth, B., Burton, D., 2008. Nitrous oxide emissions from denitrification and the partitioning of gaseous losses as affected by nitrate and carbon addition and soil aeration. Canadian Journal of Soil Science 88, 133e143. Glass, J.B., Orphan, V.J., 2012. Trace metal requirements for microbial enzymes involved in the production and consumption of methane and nitrous oxide. Frontiers in Microbiology 3, 61. Glibert, P.M., Harrison, J., Heil, C., Seitzinger, S., 2006. Escalating worldwide use of urea - a global change contributing to coastal eutrophication. Biogeochemistry 77, 441e463. Godde, M., Conrad, R., 2000. Influence of soil properties on the turnover of nitric oxide and nitrous oxide by nitrification and denitrification at constant temperature and moisture. Biology and Fertility of Soils 32, 120e128. Godfrey, L.V., Glass, J.B., 2011. The geochemical record of the ancient nitrogen cycle, nitrogen isotopes, and metal cofactors. Methods in Enzymology 486, 483e506. Goreau, T.J., Kaplan, W.A., Wofsy, S.C., Mcelroy, M.B., Valois, F.W., Watson, S.W., 1980. Production of NO2 and N2O by nitrifying bacteria at reduced concentrations of oxygen. Applied and Environmental Microbiology 40, 526e532. Grace, P., Rowlings, D., Rochester, I., Kiese, R., Butterbach-Bahl, K., 2010. Nitrous Oxide Emissions from Irrigated Cotton Soils of Northern Australia. Grandy, A.S., Loecke, T.D., Parr, S., Robertson, G.P., 2006. Long-term trends in nitrous oxide emissions, soil nitrogen, and crop yields of till and no-till cropping systems. Journal of Environmental Quality 35, 1487e1495. Grant, R.F., Pattey, E., Goddard, T.W., Kryzanowski, L.M., Puurveen, H., 2006. Modeling the effects of fertilizer application rate on nitrous oxide emissions. Soil Science Society of America Journal 70, 235e248. Gregorich, E.G., Rochette, P., VandenBygaart, A.J., Angers, D.A., 2005. Greenhouse gas contributions of agricultural soils and potential mitigation practices in Eastern Canada. Soil and Tillage Research 83, 53e72. Gruber, N., Galloway, J.N., 2008. An Earth-system perspective of the global nitrogen cycle. Nature 451, 293e296. Grundmann, G., Rolston, D., Kachanoski, R., 1988. Field soil properties influencing the variability of denitrification gas fluxes. Soil Science Society of America Journal 52, 1351e1355. Gubry-Rangin, C., Hai, B., Quince, C., Engel, M., Thomson, B.C., James, P., Schloter, M., Griffiths, R.I., Prosser, J.I., Nicol, G.W., 2011. Niche specialization of terrestrial archaeal ammonia oxidizers. Proceedings of the National Academy of Sciences of the United States of America 108, 21206e21211. Halvorson, A.D., Bartolo, M.E., Reule, C.A., Berrada, A., 2008. Nitrogen effects on onion yield under drip and furrow irrigation. Agronomy Journal 100, 1062e1069. Halvorson, A.D., Del Grosso, S.J., 2013. Nitrogen placement and source effects on nitrous oxide emissions and yields of irrigated corn. Journal of Environmental Quality 42, 312e322. Halvorson, A.D., Del Grosso, S.J., Alluvione, F., 2010a. Nitrogen source effects on nitrous oxide emissions from irrigated no-till corn. Journal of Environmental Quality 39, 1554e1562. Halvorson, A.D., Del Grosso, S.J., Alluvione, F., 2010b. Tillage and inorganic nitrogen source effects on nitrous oxide emissions from irrigated cropping systems. Soil Science Society of America Journal 74, 436e445. Hanson, B., Bendixen, W., May, D., 2000. Patterns of soil moisture, soil salinity, and soil nitrate under drip irrigation of row crops. In: Proceedings of the 2000 National Conference and Exhibition: WatereEssential for Life. Irrigation Association of Australia, Melbourne, AU. Hanson, B., May, D., 2007. The effect of drip line placement on yield and quality of dripirrigated processing tomatoes. Irrigation and Drainage Systems 21, 109e118.

172


Hao, X., Chang, C., Carefoot, J., Janzen, H., Ellert, B., 2001. Nitrous oxide emissions from an irrigated soil as affected by fertilizer and straw management. Nutrient Cycling in Agroecosystems 60, 1e8. Harrison, R.M., Yamulki, S., Goulding, K.W.T., Webster, C.P., 1995. Effect of fertilizer application on NO and N2O fluxes from agricultural fields. Journal of Geophysical Research: Atmospheres 100, 25923e25931. Hatfield, J.L., Parkin, T.B., 2014. Enhanced efficiency fertilizers: effect on agronomic performance of corn in Iowa. Agronomy Journal 106, 771e780. Hauck, R., Stephenson, H., 1965. Nitrogen sources, nitrification of nitrogen fertilizers. Effect of nitrogen source, size and pH of granule, and concentration. Journal of Agricultural and Food Chemistry 13, 486e492. Hawkins, S., Robinson, K., Layton, A., Sayler, G., 2010. Limited impact of free ammonia on Nitrobacter spp. inhibition assessed by chemical and molecular techniques. Bioresource Technology 101, 4513e4519. Heil, J., Liu, S., Vereecken, H., Bru¨ggemann, N., 2015. Abiotic nitrous oxide production from hydroxylamine in soils and their dependence on soil properties. Soil Biology and Biochemistry 84, 107e115. Heńault, C., Grossel, A., Mary, B., Roussel, M., Leónard, J., 2012. Nitrous oxide emission by agricultural soils: a review of spatial and temporal variability for mitigation. Pedosphere 22, 426e433. Hoben, J.P., Gehl, R.J., Millar, N., Grace, P.R., Robertson, G.P., 2011. Nonlinear nitrous oxide (N2O) response to nitrogen fertilizer in on-farm corn crops of the US Midwest. Global Change Biol 17, 1140e1152. Hochstein, L.I., Betlach, M., Kritikos, G., 1984. The effect of oxygen on denitrification during steady-state growth of paracoccus-halodenitrificans. Archives of Microbiology 137, 74e78. Hochstein, L.I., Tomlinson, G.A., 1988. The enzymes associated with denitrification. Annual Reviews in Microbiology 42, 231e261. Hollocher, T.C., Tate, M.E., Nicholas, D.J.D., 1981. Oxidation of ammonia by nitrosomonaseuropaea - definitive O-18-Tracer evidence that hydroxylamine formation involves a mono-oxygenase. Journal of Biological Chemistry 256, 834e836. Honda, N., Hirai, M., Ano, T., Shoda, M., 1998. Antifungal effect of a heterotrophic nitrifier Alcaligenes faecalis. Biotechnology Letters 20, 703e705. Hooper, A.B., 1968. A nitrite-reducing enzyme from Nitrosomonas europaea - preliminary characterization with hydroxylamine as electron donor. Biochimica et Biophysica Acta 162, 49. Hooper, A.B., Terry, K.R., 1979. Hydroxylamine oxidoreductase of Nitrosomonas production of nitric-oxide from hydroxylamine. Biochimica et Biophysica Acta 571, 12e20. Hultgreen, G., 2003. The Effect of Nitrogen Fertilizer Placement, Formulation, Timing, and Rate on Greenhouse Gas Emissions and Agronomic Performance. Huygens, D., Boeckx, P., Templer, P., Paulino, L., Van Cleemput, O., Oyarzuń, C., Mu¨ller, C., Godoy, R., 2008. Mechanisms for retention of bioavailable nitrogen in volcanic rainforest soils. Nature Geoscience 1, 543e548. Hyatt, C.R., Venterea, R.T., Rosen, C.J., McNearney, M., Wilson, M.L., Dolan, M.S., 2010. Polymer-coated urea maintains potato yields and reduces nitrous oxide emissions in a Minnesota loamy sand. Soil Science Society of America Journal 74, 419e428. Hynes, R.K., Knowles, R., 1982. Effect of acetylene on autotrophic and heterotrophic nitrification. Canadian Journal of Microbiology 28, 334e340. IPCC, 2007a. Climate change 2007: the physical science basis. Agenda 6, 333. IPCC, 2007b. Mitigation of climate change. In: Metz, B., Davidson, O.R., Bosch, P.R., Dave, R., Meyer, L.A. (Eds.), Contribution of Working Group III to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge Univ Press, Cambridge, UK. Climate Change 2007. IPCC, 2007c. Working group I: the physical science basis. In: Solomon, S., Qin, D., Manning, M., Chen, Z., Marquis, M., Averyt, K.B., Tignor, M., Miler, H.L. (Eds.), IPCC Fourth Assessment Report: Climate Change 2007. Cambridge University Press, Cambridge, UK.

References 173 IPCC, 2013. Anthropogenic and natural radiative forcing. In: Stocker, T.F., Qin, D., Plattner, G.-K., Tignor, M., Allen, S.K., Boschung, J., Nauels, A., Xia, Y., Bex, V., Midgley, P.M. (Eds.), Climate Change 2013: The Physical Science Basis. Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA. IPCC, 2014. In: Edenhofer, O., Pichs-Madruga, R., Sokona, Y., Farahani, E., Kadner, S., Seyboth, K., Adler, A., Baum, I., Brunner, S., Eickemeier, P., Kriemann, B., Savolainen, J., Schlo¨ mer, S., von Stechow, C., Zwickel, T., Minx, J.C. (Eds.), Climate Change 2014: Mitigation of Climate Change. Contribution of Working Group III to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA. Islam, A., Chen, D., White, R., 2007. Heterotrophic and autotrophic nitrification in two acid pasture soils. Soil Biology and Biochemistry 39, 972e975. Jiang, X., Hou, X., Zhou, X., Xin, X., Wright, A., Jia, Z., 2015. pH regulates key players of nitrification in paddy soils. Soil Biology and Biochemistry 81, 9e16. Kallenbach, C.M., Rolston, D.E., Horwath, W.R., 2010. Cover cropping affects soil N2O and CO2 emissions differently depending on type of irrigation. Agriculture, Ecosystems and Environment 137, 251e260. Katayanagi, N., Furukawa, Y., Fumoto, T., Hosen, Y., 2012. Validation of the DNDC-rice model by using CH4 and N2O flux data from rice cultivated in pots under alternate wetting and drying irrigation management. Soil Science and Plant Nutrition 58, 360e372. Keeney, D., Fillery, I., Marx, G., 1979. Effect of temperature on the gaseous nitrogen products of denitrification in a silt loam soil. Soil Science Society of America Journal 43, 1124e1128. Kennedy, T.L., Suddick, E.C., Six, J., 2013. Reduced nitrous oxide emissions and increased yields in California tomato cropping systems under drip irrigation and fertigation. Agriculture, Ecosystems and Environment 170, 16e27. Kessavalou, A., Mosier, A.R., Doran, J.W., Drijber, R.A., Lyon, D.J., Heinemeyer, O., 1998. Fluxes of carbon dioxide, nitrous oxide, and methane in grass sod and winter wheat-fallow tillage management. Journal of Environmental Quality 27, 1094e1104. Kessel, C., Venterea, R., Six, J., Adviento-Borbe, M.A., Linquist, B., Groenigen, K.J., 2013. Climate, duration, and N placement determine N2O emissions in reduced tillage systems: a meta-analysis. Global Change Biology 19, 33e44. Khdyer, I.I., Cho, C.M., 1983. Nitrification and denitrification of nitrogen fertilizers in a soil column. Soil Science Society of America Journal 47, 1134e1139. Kieft, T.L., Soroker, E., Firestone, M.K., 1987. Microbial biomass response to a rapid increase in water potential when dry soil is wetted. Soil Biology and Biochemistry 19, 119e126 [0038-0717]. Knowles, R., 1982. Denitrification. Microbiological Reviews 46, 43e70. Kong, A.Y., Fonte, S.J., van Kessel, C., Six, J., 2009. Transitioning from standard to minimum tillage: trade-offs between soil organic matter stabilization, nitrous oxide emissions, and N availability in irrigated cropping systems. Soil and Tillage Research 104, 256e262. Kool, D.M., Dolfing, J., Wrage, N., Van Groenigen, J.W., 2011. Nitrifier denitrification as a distinct and significant source of nitrous oxide from soil. Soil Biology and Biochemistry 43. Kool, D.M., Muller, C., Wrage, N., Oenema, O., Van Groenigen, J.W., 2009a. Oxygen exchange between nitrogen oxides and H2O can occur during nitrifier pathways. Soil Biology and Biochemistry 41, 1632e1641. Kool, D.M., Wrage, N., Oenema, O., Harris, D., Van Groenigen, J.W., 2009b. The O-18 signature of biogenic nitrous oxide is determined by O exchange with water. Rapid Communications in Mass Spectrometry 23, 104e108.

174


Kool, D.M., Wrage, N., Zechmeister-Boltenstern, S., Pfeffer, M., Brus, D., Oenema, O., Van Groenigen, J.W., 2010. Nitrifier denitrification can be a source of N2O from soil: a revised approach to the dual-isotope labelling method. European Journal of Soil Science 61, 759e772. Kreitinger, J.P., Klein, T., Novick, N., Alexander, M., 1985. Nitrification and characteristics of nitrifying microorganisms in an acid forest soil. Soil Science Society of America Journal 49, 1407e1410. Kuai, L.P., Verstraete, W., 1998. Ammonium removal by the oxygen-limited autotrophic nitrification-denitrification system. Applied and Environmental Microbiology 64, 4500e4506. Kyveryga, P.M., Blackmer, A.M., Ellsworth, J.W., Isla, R., 2004. Soil pH effects on nitrification of fall-applied anhydrous ammonia. Soil Science Society of America Journal 68, 545e551. Laughlin, R.J., Stevens, R.J., 2002. Evidence for fungal dominance of denitrification and codenitrification in a grassland soil. Soil Science Society of America Journal 66, 1540e1548. Lee, J., De Gryze, S., Six, J., 2008. Effect of Climate Change on Field Crop Production in the Central Valley of California. California Energy Commission, Sacramento, California. Lee, J., Hopmans, J.W., van Kessel, C., King, A.P., Evatt, K.J., Louie, D., Rolston, D.E., Six, J., 2009. Tillage and seasonal emissions of CO2, N2O and NO across a seed bed and at the field scale in a Mediterranean climate. Agriculture, Ecosystems and Environment 129, 378e390. Lee, J., Six, J., King, A.P., Van Kessel, C., Rolston, D.E., 2006. Tillage and field scale controls on greenhouse gas emissions. Journal of Environmental Quality 35, 714e725. Lees, H., Simpson, J., 1957. The biochemistry of the nitrifying organisms. 5. Nitrite oxidation by Nitrobacter. Biochemical Journal 65, 297. Leffelaar, P.A., 1986. Dynamics of partial anaerobiosis, denitrification, and water in a soil aggregate - experimental. Soil Science 142, 352e366. Leininger, S., Urich, T., Schloter, M., Schwark, L., Qi, J., Nicol, G.W., Prosser, J.I., Schuster, S.C., Schleper, C., 2006. Archaea predominate among ammonia-oxidizing prokaryotes in soils. Nature 442, 806e809. Leroy, B.L.M., Herath, H.M.S.K., Sleutel, S., De Neve, S., Gabriels, D., Reheul, D., Moens, M., 2008. The quality of exogenous organic matter: short-term effects on soil physical properties and soil organic matter fractions. Soil Use and Management 24, 139e147. Li, L., Pan, Y., Wu, Q., Zhou, X., Li, Z., 1988. Investigation on Amorphous ferric oxide acting as an electron acceptor in the oxidation of NH4þ under anaerobic condition. Acta Pedologica Sinica 25, 184e190. Linn, D.M., Doran, J.W., 1984. Effect of water filled pore space on carbon dioxide and nitrous oxide production in tilled and nontilled soils. Soil Science Society of America Journal 48, 1267e1272. Lipschultz, F., Zafiriou, O.C., Wofsy, S.C., Mcelroy, M.B., Valois, F.W., Watson, S.W., 1981. Production of NO and N2O by soil nitrifying bacteria. Nature 294, 641e643. Liu, S., Vereecken, H., Bru¨ggemann, N., 2014. A highly sensitive method for the determination of hydroxylamine in soils. Geoderma 232, 117e122. Liu, X.J., Mosier, A.R., Halvorson, A.D., Zhang, F.S., 2006. The impact of nitrogen placement and tillage on NO, N2O, CH4 and CO2 fluxes from a clay loam soil. Plant and Soil 280, 177e188. Lobell, D.B., Field, C.B., 2007. Global scale climateecrop yield relationships and the impacts of recent warming. Environmental Research Letters 2, 014002. Lobell, D.B., Schlenker, W., Costa-Roberts, J., 2011. Climate trends and global crop production since 1980. Science 333, 616e620. Lu, X., Cheng, G., 2009. Climate change effects on soil carbon dynamics and greenhouse gas emissions in Abies fabri forest of subalpine, southwest China. Soil Biology and Biochemistry 41, 1015e1021.

References 175 Ma, B., Wu, T., Tremblay, N., Deen, W., Morrison, M., McLaughlin, N., Gregorich, E., Stewart, G., 2010. Nitrous oxide fluxes from corn fields: on-farm assessment of the amount and timing of nitrogen fertilizer. Global Change Biology 16, 156e170. Maag, M., Vinther, F.P., 1996. Nitrous oxide emission by nitrification and denitrification in different soil types and at different soil moisture contents and temperatures. Applied Soil Ecology 4, 5e14. MacDonald, J.M., 2009. Manure Use for Fertilizer and for Energy: Report to Congress. DIANE Publishing. Maharjan, B., Venterea, R.T., 2013. Nitrite intensity explains N management effects on N2O emissions in maize. Soil Biology and Biochemistry 66, 229e238. Malhi, S., Nyborg, M., Beauchamp, E., 1994. Large granules, nests or bands: methods of increasing efficiency of fall-applied urea for small cereal grains in North America. Fertilizer Research 38, 61e87. Mann, P., Quastel, J., 1946. Manganese metabolism in soils. Nature 158, 154e156. Matson, P.A., Naylor, R., Ortiz-Monasterio, I., 1998. Integration of environmental, agronomic, and economic aspects of fertilizer management. Science 280, 112e115. Matthiessen, J.N., Kirkegaard, J.A., 2006. Biofumigation and enhanced biodegradation: opportunity and challenge in soilborne pest and disease management. Critical Reviews in Plant Sciences 25, 235e265. McSwiney, C.P., Robertson, G.P., 2005. Nonlinear response of N2O flux to incremental fertilizer addition in a continuous maize (Zea mays L.) cropping system. Global Change Biology 11, 1712e1719. Meijide, A., Diez, J.A., Sanchez-Martin, L., Lopez-Fernandez, S., Vallejo, A., 2007. Nitrogen oxide emissions from an irrigated maize crop amended with treated pig slurries and composts in a Mediterranean climate. Agriculture, Ecosystems and Environment 121, 383e394. Meiklejohn, J., 1952. Iron and the nitrifying bacteria. Journal of General Microbiology 7, R5. Menyailo, O.V., Hungate, B.A., 2006. Tree species and moisture effects on soil sources of N2O: quantifying contributions from nitrification and denitrification with O-18 isotopes. Journal of Geophysical Research-Biogeosciences 111. Metzger, M., Rounsevell, M., Acosta-Michlik, L., Leemans, R., Schro¨ter, D., 2006. The vulnerability of ecosystem services to land use change. Agriculture, Ecosystems and Environment 114, 69e85. Minami, K., Fukushi, S., 1986. Emission of nitrous oxide from a well-aerated andosol treated with nitrite and hydroxylamine. Soil Science and Plant Nutrition 32, 233e237. Montzka, S.A., Dlugokencky, E.J., Butler, J.H., 2011. Non-CO2 greenhouse gases and climate change. Nature 476, 43e50. Morton, J.F., 2007. The impact of climate change on smallholder and subsistence agriculture. Proceedings of the National Academy of Sciences of the United States of America 104, 19680e19685. Mosier, A.R., Halvorson, A.D., Reule, C.A., Liu, X.J., 2006. Net global warming potential and greenhouse gas intensity in irrigated cropping systems in northeastern Colorado. Journal of Environmental Quality 35, 1584e1598. Muller, E.B., Stouthamer, A.H., Vanverseveld, H.W., 1995. Simultaneous NH3 oxidation and N-2 production at reduced O-2 tensions by sewage-sludge subcultured with chemolithotrophic medium. Biodegradation 6, 339e349. Mulvaney, R.L., Khan, S.A., Mulvaney, C.S., 1997. Nitrogen fertilizers promote denitrification. Biology and Fertility of Soils 24, 211e220. Munkholm, L.J., Heck, R.J., Deen, B., Zidar, T., 2016. Relationship between soil aggregate strength, shape and porosity for soils under different long-term management. Geoderma 268, 52e59. Munoz, F., Mylavarapu, R., Hutchinson, C., 2005. Environmentally responsible potato production systems: a review. Journal of Plant Nutrition 28, 1287e1309. Murray, B.C., Baker, J.S., 2011. An output-based intensity approach for crediting greenhouse gas mitigation in agriculture: explanation and policy implications. Greenhouse Gas Measurement and Management 1, 27e36.

176


Myers, S.S., Zanobetti, A., Kloog, I., Huybers, P., Leakey, A.D., Bloom, A., Carlisle, E., Dietterich, L.H., Fitzgerald, G., Hasegawa, T., 2014. Rising CO2 threatens human nutrition. Nature 510, 139. Myrold, D.D., Tiedje, J.M., 1985. Establishment of denitrification capacity in soil: effects of carbon, nitrate and moisture. Soil Biology and Biochemistry 17, 819e822. Na¨gele, W., Conrad, R., 1990. Influence of soil pH on the nitrate-reducing microbial populations and their potential to reduce nitrate to NO and N2O. FEMS Microbiology Letters 74, 49e57. Nelson, D.W., Bremner, J.M., 1970. Role of soil minerals and metallic cations in nitrite decomposition and chemodenitrification in soils. Soil Biology and Biochemistry 2, 1e8. Nelson, S.D., Terry, R.E., 1996. The effects of soil physical properties and irrigation method on denitrification. Soil Science 161, 242e249. No¨mmik, H., 1956. Investigations on denitrification in soil. Acta Agriculturae Scandinavica 6, 195e228. Norman, R., Kurtz, L., Stevenson, F., 1987. Solubilization of soil organic matter by liquid anhydrous ammonia. Soil Science Society of America Journal 51, 809e812. Oenema, O., Velthof, G., Kuikman, P., 2001. Technical and policy aspects of strategies to decrease greenhouse gas emissions from agriculture. Nutrient Cycling in Agroecosystems 60, 301e315. Ogle, S.M., Swan, A., Paustian, K., 2012. No-till management impacts on crop productivity, carbon input and soil carbon sequestration. Agriculture, Ecosystems and Environment 149, 37e49. Ottow, J., Burth-Gebauer, I., El Demerdash, M., 1985. Influence of pH and partial oxygen pressure on the N2O-N to N2 ratio of denitrification. Denitrification in the Nitrogen Cycle 101e120. Springer. Papen, H., Vonberg, R., Hinkel, I., Thoene, B., Rennenberg, H., 1989. Heterotrophic nitrification by alcaligenes-faecalis - NO2 , NO3 , N2O, and NO production in exponentially growing cultures. Applied and Environmental Microbiology 55, 2068e2072. Paraskevopoulos, K., Antonyuk, S.V., Sawers, R.G., Eady, R.R., Hasnain, S.S., 2006. Insight into catalysis of nitrous oxide reductase from high-resolution structures of resting and inhibitor-bound enzyme from Achromobacter cycloclastes. Journal of Molecular Biology 362, 55e65. Parkin, T.B., Hatfield, J.L., 2010. Influence of nitrapyrin on N2O losses from soil receiving fall-applied anhydrous ammonia. Agriculture, Ecosystems and Environment 136, 81e86. Parton, W., Scurlock, J., Ojima, D., Schimel, D., Hall, D., 1995. Impact of climate change on grassland production and soil carbon worldwide. Global Change Biology 1, 13e22. Paustian, K., Andreń, O., Janzen, H.H., Lal, R., Smith, P., Tian, G., Tiessen, H., Noordwijk, M.V., Woomer, P.L., 1997. Agricultural soils as a sink to mitigate CO2 emissions. Soil Use and Management 13, 230e244. Paustian, K., Babcock, B., Hatfield, J., Kling, C., Lal, R., McCarl, B., McLaughlin, S., Mosier, A., Post, W., Rice, C., 2004. Climate Change and Greenhouse Gas Mitigation: Challenges and Opportunities for Agriculture. CAST Task Force Report 141. Paustian, K., Elliott, E.T., Peterson, G.A., Killian, K., 1995. Modelling climate, CO2 and management impacts on soil carbon in semi-arid agroecosystems. Plant and Soil 187, 351e365. Pedersen, H., Dunkin, K.A., Firestone, M.K., 1999. The relative importance of autotrophic and heterotrophic nitrification in a conifer forest soil as measured by15N tracer and pool dilution techniques. Biogeochemistry 44, 135e150. Peters, B., Casciotti, K.L., Samarkin, V.A., Madigan, M.T., Schutte, C.A., Joye, S.B., 2014. Stable isotope analyses of NO2 , NO3 and N2O in the hypersaline ponds and soils of the McMurdo Dry Valleys, Antarctica. Geochimica et Cosmochimica Acta 135, 87e101. Phillips, R.L., 2008. Denitrification in cropping systems at sub-zero soil temperatures. A review. Agronomy for Sustainable Development 28, 87e93. Phillips, R.L., Tanaka, D.L., Archer, D.W., Hanson, J.D., 2009. Fertilizer application timing influences greenhouse gas fluxes over a growing season. Journal of Environmental Quality 38, 1569e1579.

References 177 Picardal, F., 2012. Abiotic and microbial interactions during anaerobic transformations of Fe (II) and. Frontiers in Microbiology 3. Pinto, A.J., Marcus, D.N., Ijaz, U.Z., Bautista-de lose Santos, Q.M., Dick, G.J., Raskin, L., 2016. Metagenomic evidence for the presence of comammox Nitrospira-like bacteria in a drinking water system. mSphere 1 e00054e00015. Poth, M., 1986. Dinitrogen production from nitrite by a nitrosomonas isolate. Applied and Environmental Microbiology 52, 957e959. Poth, M., Focht, D.D., 1985. N-15 kinetic-analysis of N2O production by nitrosomonaseuropaea - an examination of nitrifier denitrification. Applied and Environmental Microbiology 49, 1134e1141. Powell, S.J., Prosser, J.I., 1991. Protection of Nitrosomonas-Europaea colonizing clay-minerals from inhibition by nitrapyrin. Journal of General Microbiology 137, 1923e1929. Qin, S., Yuan, H., Hu, C., Oenema, O., 2014. Determination of potential N2O-reductase activity in soil. Soil Biology and Biochemistry 70, 205e210. Rasmussen, T., Berks, B.C., Sanders-Loehr, J., Dooley, D.M., Zumft, W.G., Thomson, A.J., 2000. The catalytic center in nitrous oxide reductase, CuZ, is a copper-sulfide cluster. Biochemistry 39, 12753e12756. Ravishankara, A.R., Daniel, J.S., Portmann, R.W., 2009. Nitrous oxide (N2O): the dominant ozone-depleting substance emitted in the 21st century. Science 326, 123e125. Reay, D.S., Davidson, E.A., Smith, K.A., Smith, P., Melillo, J.M., Dentener, F., Crutzen, P.J., 2012. Global agriculture and nitrous oxide emissions. Nature Climate Change 2, 410e416. Reidsma, P., Ewert, F., Lansink, A.O., Leemans, R., 2010. Adaptation to climate change and climate variability in European agriculture: the importance of farm level responses. European Journal of Agronomy 32, 91e102. Reilly, J., Baethgen, W., Chege, F., Van De Geijn, S., Erda, L., Iglesias, A., Kenny, G., Patterson, D., Rogasik, J., Ro¨tter, R., 1996. Agriculture in a Changing Climate: Impacts and Adaptation. Rice, C., Smith, M., 1983. Nitrification of fertilizer and mineralized ammonium in no-till and plowed soil. Soil Science Society of America Journal 47, 1125e1129. Richardson, D.J., Wehrfritz, J.M., Keech, A., Crossman, L.C., Roldan, M.D., Sears, H.J., Butler, C.S., Reilly, A., Moir, J.W.B., Berks, B.C., Ferguson, S.J., Thomson, A.J., Spiro, S., 1998. The diversity of redox proteins involved in bacterial heterotrophic nitrification and aerobic denitrification. Biochemical Society Transactions 26, 401e408. Ritchie, G.A.F., Nicholas, D.J., 1972. Identification of sources of nitrous-oxide produced by oxidative and reductive processes in nitrosomonas-europaea. Biochemical Journal 126, 1181. Robertson, G.P., Tiedje, J.M., 1987. Nitrous-oxide sources in aerobic soils - nitrification, denitrification and other biological processes. Soil Biology and Biochemistry 19, 187e193. Robertson, L.A., Cornelisse, R., Devos, P., Hadioetomo, R., Kuenen, J.G., 1989. Aerobic denitrification in various heterotrophic nitrifiers. Antonie van Leeuwenhoek Journal of Microbiology 56, 289e299. Robertson, L.A., Kuenen, J.G., 1990. Combined heterotrophic nitrification and aerobic denitrification in Thiosphaera-Pantotropha and other bacteria. Antonie van Leeuwenhoek International Journal of General and Molecular Microbiology 57, 139e152. Rochette, P., 2008. No-till only increases N2O emissions in poorly-aerated soils. Soil and Tillage Research 101, 97e100. Rochette, P., van Bochove, E., Pre´vost, D., Angers, D.A., Bertrand, N., 2000. Soil carbon and nitrogen dynamics following application of pig slurry for the 19th consecutive year II. Nitrous oxide fluxes and mineral nitrogen. Soil Science Society of America Journal 64, 1396e1403. Rodriguez, V., de los Angeles Valdez-Perez, M., Luna-Guido, M., Ceballos-Ramirez, J.M., Franco-Hernańdez, O., Van Cleemput, O., Marsch, R., Thalasso, F., Dendooven, L., 2011. Emission of nitrous oxide and carbon dioxide and dynamics of mineral N in wastewater sludge, vermicompost or inorganic fertilizer amended soil at different water contents: a laboratory study. Applied Soil Ecology 49, 263e267.

178


Rudaz, A.O., Davidson, E.A., Firestone, M.K., 1991. Sources of nitrous-oxide production following wetting of dry soil. FEMS Microbiology Ecology 85, 117e124. Ru¨tting, T., Boeckx, P., Mu¨ller, C., Klemedtsson, L., 2011. Assessment of the importance of dissimilatory nitrate reduction to ammonium for the terrestrial nitrogen cycle. Biogeosciences 8, 1779e1791. Ryden, J., 1983. Denitrification loss from a grassland soil in the field receiving different rates of nitrogen as ammonium nitrate. Journal of Soil Science 34, 355e365. Saggar, S., Bolan, N., Bhandral, R., Hedley, C., Luo, J., 2004. A review of emissions of methane, ammonia, and nitrous oxide from animal excreta deposition and farm effluent application in grazed pastures. New Zealand Journal of Agricultural Research 47, 513e544. Sainju, U.M., Schomberg, H.H., Singh, B.P., Whitehead, W.F., Tillman, P.G., LachnichtWeyers, S.L., 2007. Cover crop effect on soil carbon fractions under conservation tillage cotton. Soil and Tillage Research 96, 205e218. Samarkin, V.A., Madigan, M.T., Bowles, M.W., Casciotti, K.L., Priscu, J.C., McKay, C.P., Joye, S.B., 2010. Abiotic nitrous oxide emission from the hypersaline Don Juan Pond in Antarctica. Nature Geoscience 3, 341e344. Schilt, A., Baumgartner, M., Blunier, T., Schwander, J., Spahni, R., Fischer, H., Stocker, T.F., 2010. Glacialeinterglacial and millennial-scale variations in the atmospheric nitrous oxide concentration during the last 800,000 years. Quaternary Science Reviews 29, 182e192. Schimel, J.P., Firestone, M.K., Killham, K.S., 1984. Identification of heterotrophic nitrification in a Sierran forest soil. Applied and Environmental Microbiology 48, 802e806. Schreiber, F., Wunderlin, P., Udert, K.M., Wells, G.F., 2012. Nitric oxide and nitrous oxide turnover in natural and engineered microbial communities: biological pathways, chemical reactions, and novel technologies. Frontiers in Microbiology 3. Schro¨ter, D., Cramer, W., Leemans, R., Prentice, I.C., Arau´jo, M.B., Arnell, N.W., Bondeau, A., Bugmann, H., Carter, T.R., Gracia, C.A., 2005. Ecosystem service supply and vulnerability to global change in Europe. Science 310, 1333e1337. Schuster, M., Conrad, R., 1992. Metabolism of nitric-oxide and nitrous-oxide during nitrification and denitrification in soil at different incubation conditions. FEMS Microbiology Ecology 101, 133e143. Seiter, S., Horwath, W.R., 2004. Strategies for Managing Soil Organic Matter to Supply Plant Nutrients. Soil Organic Matter in Sustainable Agriculture. CRC Press, Boca Raton, FL, pp. 269e294. Shaviv, A., 2001. Advances in controlled-release fertilizers. Advances in Agronomy 71, 1e49. Shaw, L.J., Nicol, G.W., Smith, Z., Fear, J., Prosser, J.I., Baggs, E.M., 2006. Nitrosospira spp. can produce nitrous oxide via a nitrifier denitrification pathway. Environmental Microbiology 8, 214e222. Shcherbak, J., 2014. Global meta analysis of the nonlinear response of soil nitrous oxide emissions to fertilizer nitrogen. Proceedings of the National Academy of Sciences of the United States of America 111. Sheehy, J., Regina, K., Alakukku, L., Six, J., 2015. Impact of no-till and reduced tillage on aggregation and aggregate-associated carbon in Northern European agroecosystems. Soil and Tillage Research 150, 107e113. Shipley, P.R., Messinger, J., Decker, A., 1992. Conserving residual corn fertilizer nitrogen with winter cover crops. Agronomy Journal 84, 869e876. Shoun, H., Fushinobu, S., Jiang, L., Kim, S.-W., Wakagi, T., 2012. Fungal denitrification and nitric oxide reductase cytochrome P450nor. Philosophical Transactions of the Royal Society of London B: Biological Sciences 367, 1186e1194. Shoun, H., Kim, D.-H., Uchiyama, H., Sugiyama, J., 1992. Denitrification by fungi. FEMS Microbiology Letters 94, 277e281.

References 179 Shrestha, J., Rich, J.J., Ehrenfeld, J.G., Jaffe, P.R., 2009. Oxidation of ammonium to nitrite under iron-reducing conditions in wetland soils laboratory, field demonstrations, and push-pull rate determination. Soil Sci 174, 156e164. Silver, W.L., Herman, D.J., Firestone, M.K., 2001. Dissimilatory nitrate reduction to ammonium in upland tropical forest soils. Ecology 82, 2410e2416. Simm, R., Mavinic, D., Ramey, W., 2006. A targeted study on possible free ammonia inhibition of Nitrospira. Journal of Environmental Engineering and Science 5, 365e376. Singh, B.K., Bardgett, R.D., Smith, P., Reay, D.S., 2010. Microorganisms and climate change: terrestrial feedbacks and mitigation options. Nature Reviews Microbiology 8, 779e790. Six, J., Ogle, S.M., Conant, R.T., Mosier, A.R., Paustian, K., 2004. The potential to mitigate global warming with no-tillage management is only realized when practised in the long term. Global Change Biology 10, 155e160. Skiba, U., Smith, K.A., 2000. The control of nitrous oxide emissions from agricultural and natural soils. Chemosphere - Global Change Science 2, 379e386. Smart, D.R., Alsina, M.M., Wolff, M.W., Matiasek, M.G., Schellenberg, D.L., Edstrom, J.P., Brown, P.H., Scow, K.M., 2011. N2O Emissions and Water Management in California Perennial Crops. Understanding Greenhouse Gas Emissions from Agricultural Management. American Chemical Society, Baltimore, MD, USA, pp. 227e255. Smid, A., Beauchamp, E., 1976. Effects of temperature and organic matter on denitrification in soil. Canadian Journal of Soil Science 56, 385e391. Smith, J., Wagner-Riddle, C., Dunfield, K., 2010. Season and management related changes in the diversity of nitrifying and denitrifying bacteria over winter and spring. Applied Soil Ecology 44, 138e146. Smith, K.A., McTaggart, I.P., Dobbie, K.E., Conen, F., 1998. Emissions of N2O from Scottish agricultural soils, as a function of fertilizer N. Nutrient Cycling in Agroecosystems 52, 123e130. þ Smith, M.S., 1982. Dissimilatory reduction of NO 2 to NH4 and N2O by a soil Citrobacter sp. Applied and Environmental Microbiology 43, 854e860. Smith, M.S., Tiedje, J.M., 1979. Phases of denitrification following oxygen depletion in soil. Soil Biology and Biochemistry 11, 261e267. Smith, P., Clark, H., Dong, H., Elsiddig, E., Haberl, H., Harper, R., House, J., Jafari, M., Masera, O., Mbow, C., 2014. Agriculture, Forestry and Other Land Use (AFOLU). Smith, P., Martino, D., Cai, Z., Gwary, D., Janzen, H., Kumar, P., McCarl, B., Ogle, S., O’Mara, F., Rice, C., 2008. Greenhouse gas mitigation in agriculture. Philosophical Transactions of the Royal Society of London B: Biological Sciences 363, 789e813. Smukler, S., O’Geen, A., Jackson, L., 2012. Assessment of best management practices for nutrient cycling: a case study on an organic farm in a Mediterranean-type climate. Journal of Soil and Water Conservation 67, 16e31. Snapp, S., Swinton, S., Labarta, R., Mutch, D., Black, J., Leep, R., Nyiraneza, J., O’Neil, K., 2005. Evaluating cover crops for benefits, costs and performance within cropping system niches. Agronomy Journal 97, 322e332. Snyder, C.S., Bruulsema, T.W., Jensen, T.L., Fixen, P.E., 2009. Review of greenhouse gas emissions from crop production systems and fertilizer management effects. Agriculture, Ecosystems and Environment 133, 247e266. Sørensen, J., Thorling, L., 1991. Stimulation by lepidocrocite (7-FeOOH) of Fe (II)-dependent nitrite reduction. Geochimica et Cosmochimica Acta 55, 1289e1294. Stanford, G., Legg, J., Dzienia, S., Simpson Jr., E., 1975. Denitrification and associated nitrogen transformations in soils. Soil Science 120, 147e152. Stark, J.M., Firestone, M.K., 1995. Mechanisms for soil-moisture effects on activity of nitrifying bacteria. Applied and Environmental Microbiology 61, 218e221. Steenwerth, K., Belina, K., 2008. Cover crops and cultivation: impacts on soil N dynamics and microbiological function in a Mediterranean vineyard agroecosystem. Applied Soil Ecology 40, 370e380.

180


Steenwerth, K.L., Hodson, A.K., Bloom, A.J., Carter, M.R., Cattaneo, A., Chartres, C.J., Hatfield, J.L., Henry, K., Hopmans, J.W., Horwath, W.R., 2014. Climate-smart agriculture global research agenda: scientific basis for action. Agriculture and Food Security 3, 1. Stehfest, E., Bouwman, L., 2006. N2O and NO emission from agricultural fields and soils under natural vegetation: summarizing available measurement data and modeling of global annual emissions. Nutrient Cycling in Agroecosystems 74, 207e228. Stein, L.Y., 2011. Surveying N2O-Producing pathways in bacteria. Methods in Enzymology: Research on Nitrification and Related Processes 486 (Part A), 131e152. Stein, L.Y., Yung, Y.L., 2003. Production, isotopic composition, and atmospheric fate of biologically produced nitrous oxide. Annual Review of Earth and Planetary Sciences 31, 329e356. Stevens, R., Laughlin, R., 1998. Measurement of nitrous oxide and di-nitrogen emissions from agricultural soils. Nutrient Cycling in Agroecosystems 52, 131e139. Stevens, R.J., Laughlin, R.J., Malone, J.P., 1998. Soil pH affects the processes reducing nitrate to nitrous oxide and di-nitrogen. Soil Biology and Biochemistry 30, 1119e1126. Stevenson, F., Harrison, R., Wetselaar, R., Leeper, R., 1970. Nitrosation of soil organic matter: III. Nature of gases produced by reaction of nitrite with lignins, humic substances, and phenolic constituents under neutral and slightly acidic conditions. Soil Science Society of America Journal 34, 430e435. Stevenson, F., Swaby, R., 1964. Nitrosation of soil organic matter: I. Nature of gases evolved during nitrous acid treatment of lignins and humic substances. Soil Science Society of America Journal 28, 773e778. Stieglmeier, M., Mooshammer, M., Kitzler, B., Wanek, W., Zechmeister-Boltenstern, S., Richter, A., Schleper, C., 2014. Aerobic nitrous oxide production through N-nitrosating hybrid formation in ammonia-oxidizing archaea. The ISME Journal 8, 1135e1146. Stroo, H.F., Klein, T.M., Alexander, M., 1986. Heterotrophic nitrification in an acid forest soil and by an acid-tolerant fungus. Applied and Environmental Microbiology 52, 1107e1111. Sutka, R.L., Ostrom, N.E., Ostrom, P.H., Breznak, J.A., Gandhi, H., Pitt, A.J., Li, F., 2006. Distinguishing nitrous oxide production from nitrification and denitrification on the basis of isotopomer abundances. Applied and Environmental Microbiology 72, 638e644. Syakila, A., Kroeze, C., 2011. The global nitrous oxide budget revisited. Greenhouse Gas Measurement and Management 1, 17e26. Szukics, U., Abell, G.C., Ho¨dl, V., Mitter, B., Sessitsch, A., Hackl, E., ZechmeisterBoltenstern, S., 2010. Nitrifiers and denitrifiers respond rapidly to changed moisture and increasing temperature in a pristine forest soil. FEMS Microbiology Ecology 72, 395e406. Taub, D.R., Miller, B., Allen, H., 2008. Effects of elevated CO2 on the protein concentration of food crops: a meta-analysis. Global Change Biology 14, 565e575. Tenuta, M., Beauchamp, E.G., 2003. Nitrous oxide production from granular nitrogen fertilizers applied to a silt loam. Canadian Journal of Soil Science 83, 521e532. Thauer, R.K., Jungermann, K., Decker, K., 1977. Energy conservation in chemotropic anaerobic bacteria. Bacteriological Reviews 41, 100e180. Thomson, A.J., Giannopoulos, G., Pretty, J., Baggs, E.M., Richardson, D.J., 2012. Biological sources and sinks of nitrous oxide and strategies to mitigate emissions. Philosophical Transactions of the Royal Society of London B: Biological Sciences 367, 1157e1168. Thorn, K., Mikita, M., 2000. Nitrite fixation by humic substances nitrogen-15 nuclear magnetic resonance evidence for potential intermediates in chemodenitrification. Soil Science Society of America Journal 64, 568e582. Thornton, F.C., Bock, B.R., Tyler, D.D., 1996. Soil emissions of nitric oxide and nitrous oxide from injected anhydrous ammonium and urea. Journal of Environmental Quality 25, 1378e1384. Tiedje, J.M., 1988. Ecology of denitrification and dissimilatory nitrate reduction to ammonia. In: Zehnder, A.J.B. (Ed.), Biology of Anaerobic Microorganisms, vol. 66.

References 181 Tiedje, J.M., Sexstone, A.J., Myrold, D.D., Robinson, J.A., 1983. Denitrification: ecological niches, competition and survival. Antonie van Leeuwenhoek 48, 569e583. Tiedje, J.M., Sexstone, A.J., Parkin, T.B., Revsbech, N.P., Shelton, D.R., 1984. Anaerobic processes in soil. Plant and Soil 76, 197e212. Tourna, M., Stieglmeier, M., Spang, A., Ko¨nneke, M., Schintlmeister, A., Urich, T., Engel, M., Schloter, M., Wagner, M., Richter, A., 2011. Nitrososphaera viennensis, an ammonia oxidizing archaeon from soil. Proceedings of the National Academy of Sciences of the United States of America 108, 8420e8425. Trenkel, M.E., 2010. Slow-and Controlled-release and Stabilized Fertilizers: An Option for Enhancing Nutrient Use Efficiency in Agriculture. IFA, International fertilizer industry association. Turk, O., Mavinic, D.S., 1989. Maintaining nitrite build-up in a system acclimated to free ammonia. Water Research 23, 1383e1388. USDA, 2010. Agricultural Resource Management Survey. U.S. Department of Agriculture, Economic Research Service, Washington, DC (verified July 23, 2016). http://www.ers. usda.gov/Data/ARMS/. USDA, 2014. Quantifying Greenhouse Gas Fluxes in Agriculture and Forestry: Methods for Entity-Scale Inventory. USEPA, 2009. Inventory of U.S. Greenhouse Gas Emissions and Sinks: 1990-2007. http://epa. gov/climatechange/emissions/usinventoryreport.html. USEPA, 2016. Inventory of U.S. Greenhouse Gas Emissions and Sinks: 1990-2014. U.S. Environmental Protection Agency, 1200 Pennsylvania Ave. N.W. Washington, DC 20460. EPA 430-R-10-006 (verified May 27, 2016). http://epa.gov/climatechange/ emissions/usinventoryreport.html. Van Cleemput, O., 1998. Subsoils: chemo- and biological denitrification, N2O and N-2 emissions. Nutrient Cycling in Agroecosystems 52, 187e194. Van Cleemput, O., Baert, L., 1984. Nitrite - a key compound in N-Loss processes under acid conditions. Plant and Soil 76, 233e241. Van Cleemput, O., Samater, A.H., 1996. Nitrite in soils: accumulation and role in the formation of gaseous N compounds. Fertilizer Research 45, 81e89. Van Groenigen, J.W., Velthof, G.L., Oenema, O., Van Groenigen, K.J., Van Kessel, C., 2010. Towards an agronomic assessment of N2O emissions: a case study for arable crops. European Journal of Soil Science 61, 903e913. van Kessel, M.A., Speth, D.R., Albertsen, M., Nielsen, P.H., den Camp, H.J.O., Kartal, B., Jetten, M.S., Lu¨cker, S., 2015. Complete nitrification by a single microorganism. Nature 528, 555e559. Varco, J., Frye, W., Smith, M., Grove, J., 1987. Legume N Transformation and Recovery by Corn as Influenced by Tillage. The Role of Legumes in Conservation Tillage Systems. Soil Conservation Society of America, University of Georgia, Athens. Vaughan, S.M., Dalal, R.C., Harper, S.M., Menzies, N.W., 2011. Effect of fresh green waste and green waste compost on mineral nitrogen, nitrous oxide and carbon dioxide from a Vertisol. Waste Management 31, 1720e1728. Venkiteswaran, J.J., Rosamond, M.S., Schiff, S.L., 2014. Nonlinear response of riverine N2O fluxes to oxygen and temperature. Environmental Science and Technology 48, 1566e1573. Venterea, R.T., 2007. Nitrite-driven nitrous oxide production under aerobic soil conditions: kinetics and biochemical controls. Global Change Biology 13, 1798e1809. Venterea, R.T., Burger, M., Spokas, K.A., 2005. Nitrogen oxide and methane emissions under varying tillage and fertilizer management. Journal of Environmental Quality 34, 1467e1477. Venterea, R.T., Dolan, M.S., Ochsner, T.E., 2010. Urea decreases nitrous oxide emissions compared with anhydrous ammonia in a Minnesota corn cropping system. Soil Science Society of America Journal 74, 407e418. Venterea, R.T., Maharjan, B., Dolan, M.S., 2011. Fertilizer source and tillage effects on yieldscaled nitrous oxide emissions in a corn cropping system. Journal of Environmental Quality 40, 1521e1531.

182


Venterea, R.T., Rolston, D.E., 2000. Nitric and nitrous oxide emissions following fertilizer application to agricultural soil: biotic and abiotic mechanisms and kinetics. Journal of Geophysical Research: Atmospheres 105, 15117e15129. Venterea, R.T., Stanenas, A.J., 2008. Profile analysis and modeling of reduced tillage effects on soil nitrous oxide flux. Journal of Environmental Quality 37, 1360e1367. Vermeulen, S.J., Campbell, B.M., Ingram, J.S., 2012. Climate change and food systems. Annual Review of Environment and Resources 37, 195. Walker, C., De La Torre, J., Klotz, M., Urakawa, H., Pinel, N., Arp, D., Brochier-Armanet, C., Chain, P., Chan, P., Gollabgir, A., 2010. Nitrosopumilus maritimus genome reveals unique mechanisms for nitrification and autotrophy in globally distributed marine crenarchaea. Proceedings of the National Academy of Sciences of the United States of America 107, 8818e8823. Wang, Q., Burger, M., Doane, T., Horwath, W., Castillo, A., Mitloehner, F., 2013. Effects of inorganic v. organic copper on denitrification in agricultural soil. Advances in Animal Biosciences 4, 42e49. Watson, C., Atkinson, D., Gosling, P., Jackson, L., Rayns, F., 2002. Managing soil fertility in organic farming systems. Soil Use and Management 18, 239e247. Weber, D.F., Gainey, P., 1962. Relative sensitivity of nitrifying organisms to hyrogen ions in soils and in solutions. Soil Science 94, 138e145. Webster, E.A., Hopkins, D.W., 1996. Contributions from different microbial processes to N2O emission from soil under different moisture regimes. Biology and Fertility of Soils 22, 331e335. Wei, X.M., Vajrala, N., Hauser, L., Sayavedra-Soto, L.A., Arp, D.J., 2006. Iron nutrition and physiological responses to iron stress in Nitrosomonas europaea. Archives of Microbiology 186, 107e118. Weier, K.L., Doran, J.W., Power, J.F., Walters, D.T., 1993. Denitrification and the dinitrogen nitrous oxide ratio as affected by soil-water, available carbon, and nitrate. Soil Science Society of America Journal 57, 7. Williams, E., Hutchinson, G., Fehsenfeld, F., 1992. NOx and N2O emissions from soil. Global Biogeochemical Cycles 6, 351e388. Winogradsky, S., 1890. Recherches sur les organismes de la nitrification. Annales de l’Institut Pasteur 4, 11e12. Wolfe, D., 1997. Soil Compaction: Crop Response and Remediation. Report. Wood, P., 1990. Autotrophic and heterotrophic mechanisms for ammonia oxidation. Soil Use and Management 6, 78e79. Wrage, N., van Groenigen, J.W., Oenema, O., Baggs, E.M., 2005. A novel dual-isotope labelling method for distinguishing between soil sources of N2O. Rapid Communications in Mass Spectrometry 19, 3298e3306. Wrage, N., Velthof, G.L., Laanbroek, H.J., Oenema, O., 2004a. Nitrous oxide production in grassland soils: assessing the contribution of nitrifier denitrification. Soil Biology and Biochemistry 36, 229e236. Wrage, N., Velthof, G.L., Oenema, O., Laanbroek, H.J., 2004b. Acetylene and oxygen as inhibitors of nitrous oxide production in Nitrosomonas europaea and Nitrosospira briensis: a cautionary tale. FEMS Microbiology Ecology 47, 13e18. Wrage, N., Velthof, G.L., van Beusichem, M.L., Oenema, O., 2001. Role of nitrifier denitrification in the production of nitrous oxide. Soil Biology and Biochemistry 33, 1723e1732. Wright, A.L., Provin, T.L., Hon, F.M., Zuberer, D.A., White, R.H., 2008. Compost impacts on dissolved organic carbon and available nitrogen and phosphorus in turfgrass soil. Waste Management 28, 1057e1063. Xu, J.-Z., Peng, S.-Z., Hou, H.-J., Yang, S.-H., Luo, Y.-F., Wang, W.-G., 2013. Gaseous losses of nitrogen by ammonia volatilization and nitrous oxide emissions from rice paddies with different irrigation management. Irrigation Science 31, 983e994. Xu-Ri, Prentice, I.C., Spahni, R., Niu, H.S., 2012. Modelling terrestrial nitrous oxide emissions and implications for climate feedback. New Phytologist 196, 472e488.

References 183 Yang, W.D.H., Teh, Y.A., Silver, W.L., 2011. A test of a field-based N-15-nitrous oxide pool dilution technique to measure gross N2O production in soil. Global Change Biology 17, 3577e3588. Yang, W.H., Weber, K.A., Silver, W.L., 2012. Nitrogen loss from soil through anaerobic ammonium oxidation coupled to iron reduction. Nature Geoscience 5, 538e541. Yates, T., Si, B., Farrell, R., Pennock, D., 2006. Probability distribution and spatial dependence of nitrous oxide emission. Soil Science Society of America Journal 70, 753e762. Ye, R., Horwath, W.R., 2016. Nitrous oxide uptake in rewetted wetlands with contrasting soil organic carbon contents. Soil Biology and Biochemistry 100, 110e117. Yun, H.J., Kim, D.J., 2003. Nitrite accumulation characteristics of high strength ammonia wastewater in an autotrophic nitrifying biofilm reactor. Journal of Chemical Technology and Biotechnology 78, 377e383. Zaman, M., Matsushima, M., Chang, S.X., Inubushi, K., Nguyen, L., Goto, S., Kaneko, F., Yoneyama, T., 2004. Nitrogen mineralization, N2O production and soil microbiological properties as affected by long-term applications of sewage sludge composts. Biology and Fertility of Soils 40, 101e109. Zebarth, B., Rochette, P., Burton, D., Price, M., 2008a. Effect of fertilizer nitrogen management on N2O emissions in commercial corn fields. Canadian Journal of Soil Science 88, 189e195. Zebarth, B.J., Rochette, P., Burton, D.L., 2008b. N2O emissions from spring barley production as influenced by fertilizer nitrogen rate. Canadian Journal of Soil Science 88, 197e205. Zhang, J., Cai, Z., Zhu, T., 2011. N2O production pathways in the subtropical acid forest soils in China. Environmental research 111, 643e649. Zhang, J., Sun, W., Zhong, W., Cai, Z., 2014. The substrate is an important factor in controlling the significance of heterotrophic nitrification in acidic forest soils. Soil Biology and Biochemistry 76, 143e148. Zhang, L.-M., Hu, H.-W., Shen, J.-P., He, J.-Z., 2012. Ammonia-oxidizing archaea have more important role than ammonia-oxidizing bacteria in ammonia oxidation of strongly acidic soils. The ISME journal 6, 1032e1045. Zhou, Z., Takaya, N., Sakairi, M.A.C., Shoun, H., 2001. Oxygen requirement for denitrification by the fungus Fusarium oxysporum. Archives of Microbiology 175, 19e25. Zhu, K., Bruun, S., Larsen, M., Glud, R.N., Jensen, L.S., 2014a. Spatial oxygen distribution and nitrous oxide emissions from soil after manure application: a novel approach using planar optodes. Journal of Environmental Quality 43, 1809e1812. Zhu, X., Burger, M., Doane, T.A., Horwath, W.R., 2013a. Ammonia oxidation pathways and nitrifier denitrification are significant sources of N2O and NO under low oxygen availability. Proceedings of the National Academy of Sciences of the United States of America 110, 6328e6333. Zhu, X., Burger, M., Waterhouse, H., Horwath, W.R., 2014b. The Effect of Ammonical N Fertilizer Concentration Soil O2 Consumption and N2O Production Pathways. ASA, CSSA and SSSA International Annual Meetings. Zhu, X., Silva, L.C.R., Doane, T.A., Wu, N., Horwath, W.R., 2013b. Quantifying the effects of green waste compost application, water content and nitrogen fertilization on nitrous oxide emissions in 10 agricultural soils. Journal of Environmental Quality 42, 912e918. Zhu-Barker, X., Cavazos, A.R., Ostrom, N.E., Horwath, W.R., Glass, J.B., 2015a. The importance of abiotic reactions for nitrous oxide production. Biogeochemistry 126, 251e267. Zhu-Barker, X., Doane, T.A., Horwath, W.R., 2015b. Role of green waste compost in the production of N2O from agricultural soils. Soil Biology and Biochemistry 83. Zhu-Barker, X., Horwath, W.R., Burger, M., 2015c. Knife-injected anhydrous ammonia increases yield-scaled N2O emissions compared to broadcast or band-applied ammonium sulfate in wheat. Agriculture, Ecosystems and Environment 212, 148e157.


Chapter | Seven

The Response of Forest Ecosystems to Climate Change

Armando G omez-Guerrero*, 1, Timothy Doanex

*Colegio de Postgraduados, Postgrado en Ciencias Forestales, Carretera Me´xicoTexcoco, Montecillo, Estado de Me´xico; xDepartment of Land, Air and Water Resources, University of California, Davis, CA, United States 1


INTRODUCTION In discussing the most likely effects of climate change on the biogeochemical processes that occur in forest soils, it is important to consider the relationship among the three compartments of terrestrial ecosystems: the atmosphere (climate), the biosphere (vegetation), and the lithosphere (soil) (Campbell et al., 2009). Changes in any one of these compartments will impact the others according to net transfer of mass and energy and mean residence times established among compartments (Schlesinger and Bernhardt, 2013). The incorporation of organic matter into the soil is one of the most important processes in soil formation, and changes in net primary productivity (NPP) and quality of detritus can have a profound impact on soil processes (Breymeyer, 2003; McTiernan et al., 2003; Santonja et al., 2015). Higher temperatures can influence water stress, phenology, and the species composition of a forest, all of which will affect the mean residence time of canopy foliage. The degree of retention of foliage in turn influences internal nutrient cycles in vegetation as well as transfer of nutrients from canopy to soil. The effects of climate on the phenology and composition of forest species can therefore indirectly impact soil processes by modifying the amount and quality of the litter. Similarly, any changes in climate on litter quality and soil moisture can consequently change root activity and soil microbial activity, including the rate of decomposition of existing soil organic matter and litter. This in turn can affect fluxes of the greenhouse gases CO2, CH4, and N2O to the atmosphere, changing its capacity to retain radiant energydthe focal point of climate change (Campbell et al., 2009; Fang et al., 2014; McDaniel et al., 2014). Climate Change Impacts on Soil Processes and Ecosystem Properties. https://doi.org/10.1016/B978-0-444-63865-6.00007-7 Copyright © 2018 Elsevier B.V. All rights reserved.

185

186


From existing literature it can be inferred that the effect of climate change on plant growth will vary according to environmental conditions and plant physiological response. Predictions of this effect often differ; however, some anticipated impacts on forest ecosystems and their soils have been identified as follows: (1) increases in atmospheric CO2 will increase NPP and soil carbon, although the extent of this effect will depend on the availability of soil water and nutrients, both limiting factors in forest productivity (Chertov et al., 2010; Ge et al., 2013; Kirschbaum et al., 2012; Poulter et al., 2013; Wang et al., 2012); (2) warmer temperatures will lead to changes in soil and plant respiration and nitrogen mineralization, as well as variations in precipitation and longer growing seasons (Campbell et al., 2009; Schindlbacher et al., 2012); and (3) management practices favoring dominance of commercial tree species will cause reductions in soil carbon stocks in both natural and managed forests (Ge et al., 2013; Noormets et al., 2015). Owing to the complex nature of the interactions among climate, vegetation, and soil, the response of tree communities to CO2 increases can vary. For example, depending on soil water availability, the effect of increased atmospheric CO2 and temperature on forest growth can be positive or negative (Ge et al., 2013; Norby and Zak, 2011). This disparity in observed effects is most evident in the climatic component of an ecosystem: changes in mean annual temperature and precipitation, combined with spatially variable soil properties, result in environmental changes from microtopographic up to regional scales (Munir et al., 2014; von Arx et al., 2013; Berner et al., 2017). In addition to the major changes in climate (temperature and precipitation), other factors such as nitrogen deposition and ozone damage (Campbell et al., 2009; Grulke and Tausz, 2014) are also expected to affect NPP of forest ecosystems. The positive effect of N deposition on growth will be more evident in forest systems than in agricultural lands where an important part of the productivity is removed annually; higher availability of N by deposition leads to increased reservoirs of organic matter and aboveground biomass (Davies et al., 2016). The initial stages of increased N deposition have positive effects on NPP but advanced stages can lead to negative nutritional effects due to soil acidification and removal of soil cations (Fenn et al., 2006; Silva et al., 2015). Simulated scenarios of N saturation indicate that forest NPP will decline and N cycling will be slowed by higher N uptake in biomass and lower rates of N mineralization (Lu et al., 2016). The damage caused by ozone will be another confounding factor among the effects of climate change on vegetation. For example, defoliation of Pinus halepensis L. in the Mediterranean region of France was attributed to increased temperature (0.54 C per decade), the reduction of days with precipitation (2.5 days per decade), and lower soil moisture. Similar symptoms for Pinus cembra Mill. stands, however, were attributed to ozone damage, which is more acute at higher elevations. Owing to more favorable soil moisture conditions at higher altitudes, where P. cembra grows, longer periods of open stomata increased susceptibility to ozone damage (Sicard and Dalstein-Richier, 2015).

Global Distribution of Studies on Climate Change and Forest Soils

187

The effects of climate change on ecosystem NPP in general are reflected in soil processes via changes in the production of litter and changes in heterotrophic and autotrophic soil respiration. Research on the effects of climate change in forest ecosystems is not extensive, primarily because it is a relatively new phenomenon, and also because studies are not evenly distributed across the forest ecosystems of the world. In addition, forests grow and respond slowly to alterations in their environment, and trees lives for hundreds of years, making it difficult to determine the factors having the greatest effect on tree growth during the history of a forest. In general, the most studied forests ecosystems are the temperate and boreal forests of North America and Europe. In contrast, there are extensive regions of tropical forests where responses to climate change have not been investigated. Indeed, regarding the carbon cycle, estimates of carbon stocks of tropical ecosystems are the most uncertain (Pan et al., 2011). This chapter describes the ways in which soil processes in forests will likely respond to climate change, emphasizing the connection and interdependence between atmosphere, vegetation, and soil.

GLOBAL DISTRIBUTION OF STUDIES ON CLIMATE CHANGE AND FOREST SOILS A review of Web of Science for “climate change and forest soils” for the period 2000e16 shows that the top five countries conducting research in this topic are the United States, Germany, China, Canada, and France, representing 72% of the total studies (1342 of 1589) (Fig. 7.1). A survey of studies of “climate

FIGURE 7.1 Distribution of studies on climate change and forest soils from 2000 to 2016; these studies were carried out in countries indicated by hatched lines. Polygons in color in other countries denote types of soils according to the Food and Agriculture Organization (FAO) classification.

188


change and primary productivity of forest ecosystems” is similar, with the United Kingdom replacing France and the top five countries accounting for 73% of the total studies (180 of 247). These data demonstrate the need for studies in tropical systems, within the zone between 20oN and 20oS, for a more comprehensive understanding of the issues at hand. Prominent among tropical soils are Ferralsols (Oxisols), Acrisols (Ultisols), Lixisols (Alfisols), and Nitosols (including Oxisols, Alfisols, and Ultisols), all of which support forest vegetation (Osman, 2013). More research is needed to more accurately determine potential net changes in carbon pools as a result of climate change, including consideration of the dynamic between deforestation and forest regrowth (Friedlingstein et al., 2010). Forest regrowth in the tropics may represent a greater carbon sink than intact forests due to higher net gains above and below ground (Pan et al., 2011). Other tropical soil types that sustain forests are Cambisols (Inceptisols) and Luvisols (Alfisols). Dry environments dominated by Leptosols (Entisols) and Calcisols (Aridisols) are also important, as these ecosystems are prone to desertification, with consequent loss of their carbon reservoirs (Osman, 2013). One of the most notable tropical forests in terms of productivity is found in Borneo, Southeast Asia, where the aboveground productivity is almost 50% higher than the most productive forests in the Amazon (Banin et al., 2014). This is explained by higher annual solar radiation in Borneo, floristic structure dominated by the Dipterocarpacea family, and favorable soil properties such as high phosphorus availability and base saturation (Banin et al., 2014). This is just one example of a region where forest ecosystems are markedly different from those described in the majority of existing studies, and where additional research will prove valuable in better understanding the complete extent of potential effects of climate change. The location of experiments with forest species using enriched CO2 atmospheres to study climate effects also shows an uncoupled distribution at global scales, with very low representation of tropical and boreal forests, two very important ecosystems in terms of vegetation and soil carbon density (see Figure 1 in Jones et al., 2014). Furthermore, Jones et al. (2014) point out that reduction in funding for experimentation will prevent us from fully understanding climate effects on a global scale. Such scientific evidence of the effects of climate change is more helpful to decision-makers as more of the world’s important ecosystems, including forests, are represented.

CHANGES IN NET PRIMARY PRODUCTIVITY OF FOREST ECOSYSTEMS NPP in forest ecosystems is intimately linked to soil processes. When forest productivity increases, soil carbon more often than not also increases if the formation of stable C is promoted (Jandl et al., 2007). Higher levels of atmospheric CO2, one aspect of climate change, can stimulate productivity as trees fix more carbon, increasing not only aboveground biomass (Camarero et al., 2015) but also root growth (Norby and Zak, 2011) and root exudates (Sulman et al.,

Changes in Net Primary Productivity of Forest Ecosystems

189

2014). However, this increase in carbon fixed by trees would also increase inputs to the soil of readily assimilated forms of carbon such as sugars and amino acids, which in turn are known to accelerate the decomposition of other more stable forms of carbondthe “priming effect” (Sulman et al., 2014). The net effect of increases in NPP on carbon storage in forest soils will depend on the specific metabolic pathways affected as photosynthates travel from the leaves through the root system and into the soil (Drake et al., 2016) and the relative influence of these inputs compared to surface inputs. The response of forest ecosystems to elevated atmospheric CO2 has been studied in different ways, with direct approaches involving free-air CO2 enrichment (FACE) experiments (Duursma et al., 2016; Norby and Zak, 2011). Some FACE studies have reported higher aboveground NPP in the initial years following establishment of enriched (eCO2) conditions; this effect, however, is not sustained over longer periods (e.g., Oren et al., 2001). However, the increases in aboveground tree biomass in such experiments should not be misinterpreted or directly placed into the context of the carbon cycle because part of NPP of forest ecosystems lies in belowground components, and the effect observed in a stand of trees depends on its age (Norby and Zak, 2011; Jones et al., 2014). Younger forests have naturally higher NPP, especially in growth phases approaching canopy closure. Tree stands under eCO2 have shown increases in leaf area index (LAI) that followed a negative exponential trend across a range of LAI values ranging from 2 to 5; that is, the relative response to eCO2 is greater in stands with low LAI (Norby and Zak, 2011). Although subsequent increases in NPP per LAI unit have shown that eCO2 continues to stimulate NPP after canopy closure, this response is not yet well understood, and additional environmental factors may contribute to diminish the response over time (Norby and Zak, 2011). In contrast to these observations, the results of an ecosystem-level FACE experiment in Australia with Eucalyptus tereticornis Sm. did not show a positive response in LAI under eCO2, even at low LAI values ranging from 1.2 to 2.2 (Duursma et al., 2016). In certain cases the gain in carbon in aerial components under eCO2 can result in higher root and rhizosphere respiration with no net gain of carbon in the ecosystem as a whole (Drake et al., 2016). As pointed out before, FACE experiments have contributed importantly to understanding the response of forests to eCO2, but the actual spatial distribution of existing experiments highlights the need for more research in boreal and tropical sites (Jones et al., 2014). Using a dendrochronological series in forests in northeastern Spain, trends in basal area increment of Pinus uncinata Ram. were observed to increase in response to rising CO2, particularly in older and more mature forests (Camarero et al., 2015). Such a response was only evident in moist environments, however, lending additional support to the idea that the effect of increased atmospheric CO2 on forest NPP depends on soil water availability. Forests, like other ecosystems, are intricate networks of interrelated factors, and consequently, the overall response of a forest to climate change will depend on the influence of dominant site-specific conditions. For example, while

190


increased growth was observed in Pinus uncinata stands by Camarero et al. (2015), no response was evident for the same species after four years of eCO2 at the treeline ecotone in Central Alps in Switzerland (Handa et al., 2006). This lack of growth response was explained in terms of differences in successional stages, based on an observed 41% increase in growth of Larix decidua trees. Larix is an early successional species that can colonize undeveloped substrate on recent glacier moraines and was able to optimize use of carbon resources and weather conditions, while Pinus, a later successional species, did not grow well on undeveloped substrates even under greater availability of CO2. Furthermore, climate change will impact managed and natural forests differently, since allocation of biomass (and carbon) varies between such systems. Forest management has resulted in up to 50% reduction in carbon stocks in forests and a redistribution of carbon, with relatively more carbon present as aboveground biomass compared to belowground components; in addition, managed forests are an average of 50 years younger than natural stands, with a greater proportion of coniferous (25% more) species than deciduous species (Noormets et al., 2015). Managed forests are poised to suffer greater losses of soil carbon as a result of alterations in climate, although differences in management practices such as fertilization will lead to different outcomes. In both managed and natural forests, soil carbon will accumulate as long as gains in NPP through photosynthesis are greater than soil respiration (Jandl et al., 2007). In the case of natural forests, some effects of climate change will arise from shifts in overall tree diversity as some species are better adapted to resist changes in environmental conditions such as water availability. The composition of tree biomass is linked to nutrient dynamics. Material from different species, as a whole, decomposes faster than material from only one species. In particular, pine needles constitute a recalcitrant substrate for decomposition due to relatively high concentrations of polyphenols and terpenes, but its decomposition improves when mixed with oak leaves, providing evidence of a strong link between tree diversity, nutrient cycling, and NPP (Gueńon et al., 2017). Litter quality plays an important role in the cycling of N and P in particular, which are not simple processes that depend solely on the concentration of nutrients of the substrate, but also on the complementary presence of the community of decomposers and its affinity for the litter (Wurzburger and Hendrick, 2009; Vos et al., 2013). The effects of climate change on the way in which different tree species shape a forest are likely more complex than currently perceived. The response of forest ecosystems to increased CO2 may also depend on nutrient availability (Wieder et al., 2015), although some FACE experiments have been conducted in nitrogen-limited sites and nonetheless reported stimulation of tree growth under eCO2 (e.g., Feng et al., 2015; Norby and Zak, 2011). One explanation for this lack of relationship between soil fertility and growth response is that under eCO2, trees develop more extensive and deeper root

Changes in Net Primary Productivity of Forest Ecosystems

191

systems, exploring a greater volume of soil and therefore alleviating any nutrient deficiencies (Feng et al., 2015). Some authors have suggested that, in predicting NPP as a consequence of increased atmospheric CO2, most global models do not account for the basic stoichiometric relationship between carbon and nitrogen, leading to overestimates of biomass gains in forest ecosystems (Wang and Houlton, 2009). That nitrogen availability is a limiting factor in the productivity of forest ecosystems is well documented (Schlesinger and Bernhardt, 2013), which makes it worthwhile to further study progressive nitrogen limitation as part of understanding the relationship between climate change, NPP, and soil processes. Litter production and quality is related to nutrient availability, in that litterfall determines the amount of nutrients returned from the canopy to the forest floor and subsequent transfer to the mineral soil. Once detritus enters forest floor reservoirs, the rate of release of nutrients is determined by the activity of organisms responsible for decomposition. Compared to the litter of coniferous forests, the litter of broadleaf forests decomposes faster as it is higher in water-soluble sugars, organic acids, and amino acids (Priha et al., 2001). On the other hand, compared to coniferous wood, the wood of broadleaf species is denser, which leads to higher soil carbon stocks; in other words, at similar stand volumes broadleaf forests can accumulate more aboveground carbon (Jandl et al., 2007), which in turn may favor greater long-term storage in soil. With expected increases in temperature due to climate change, soils may lose moisture faster, influencing rates of decomposition and nutrient cycling (Santonja et al., 2015). Moreover, the quality of litter is an important factor influencing its decomposition and release of nutrients. As mentioned above, litter from multiple species decomposes faster than litter from a single species, and if climate change leads to reduced species diversity, this may indirectly affect litter decomposition (Santonja et al., 2015). At a regional scale, changes in rainfall and temperature determine the effect of climate change on forest ecosystems, while on a local scale, topography, soil properties, and soil profile characteristics also contribute to how a forest responds. Depending on its topographical position, a soil profile may vary in its ability to store water, with greater capacity for storage in deeper soils located downslope. It is predicted that the effects of climate change on pinion (Pinus edulis) and juniper (Juniperus monosperma) forests in the southern United States will depend in part on topography and soil water retention, with more negative impacts occurring on sloping terrain (Petrie et al., 2015). These authors also point out that the effects of climate change on dry forests will depend on the physiology of individual tree species, with a lower impact on droughttolerant species that respond rapidly by closing their stomata. However, the potential for increased water stress due to climate change is not limited to arid areas. In spite of the fact that the majority of rainfall across the globe occurs in tropical regions, high altitude forests (between 30 N and 30 S) show evidence of water stress due to the recent increase in average temperatures,

192


although the intensity of this effect decreases at higher altitudes (Krishnaswamy et al., 2014). Rainfall is clearly a critical factor driving NPP in arid environments, but the extent to which climate change will influence NPP in such places depends on the degree of alteration of the uniformity of precipitation throughout the growing season. Irregular distribution of rainfall in the growing season generally reduces NPP, although the strategy of individual species to resist drought events as well as the capacity of the soil to store water are also important (Xu and Wang, 2016). Simulation studies of managed Picea abies (L.) Karts forests suggest that NPP increases as temperature increases, with greater responses expected as more plant residues are incorporated into the soil to maintain nutrient cycling. Preserving soil nutrients by optimally managing tree residues could therefore enhance the positive effects of climate change on forest growth (Makipaa et al., 2015). In addition to such favorable effects on tree growth, atmospheric nitrogen deposition could stimulate NPP even further. While simulation studies have corroborated the positive effects of climate change and nitrogen deposition on forest growth in northern Europe, the additional gain in NPP ultimately depends on water availability, which can lead to contrasting effects under different otherwise similar circumstances. For example, in the Mediterranean region, the apparent benefits of nitrogen deposition and warmer temperatures would in fact result in higher defoliation, lower tree growth, and reduced carbon inputs to soil in drought-sensitive species such as Quercus ilex and Fagus sylvatica (De Marco et al., 2014). In addition to affecting tree growth and related impacts on soil, climate change will affect species distribution according to the ability to resist drought and changes in soil water and biomass allocation. After 15 years of drought treatment, plots of Quercus ilex L. and Phillyrea latifolia L. showed a reduction of 13% in soil water and an 8% reduction in net photosynthesis. Growth of Quercus trees was reduced by 40%, while only minimal effects were observed on growth of Phillyrea (Ogaya et al., 2014). The drastic reduction in growth of Quercus was attributed to differences in carbon allocationdthis species tends to prioritize allocation of carbon to belowground components and root exudates. Positive or negative changes in NPP and the resulting indirect effects on soil processes are determined on a geographic scale by changes in climate, while on a local scale these changes depend primarily on soil water availability and, to a lesser extent, by the ability of trees to resist changes in soil moisture (Davi et al., 2006). Unlike agricultural systems where water and nutrients can be adequately supplied, growth of forest species depends on the storage and supply of water and nutrients, processes which are themselves regulated by properties such as depth, organic matter content, and texture. Local topography is also important due to its influence on soil profile development. In conclusion, forest NPP can be enhanced by increased atmospheric CO2, although the actual response will ultimately depend on stand structure,

Sequestration of Carbon in Forest Soils

193

efficiency in capturing carbon, and the availability of resources, most significantly soil water. Furthermore, different species respond differently, and any resulting shifts in species distribution further contribute to the overall productivity of a forest. Changes in soil carbon in forest ecosystems will be determined by the balance between litter deposition and soil respiration, a balance which will also shift according to alterations in aboveground and belowground processes imposed by climate change.

SEQUESTRATION OF CARBON IN FOREST SOILS Globally, most organic carbon is located in two types of ecosystems: boreal forests and tropical forests, which represent 87% (743 Pg) of global stock, with a contrasting distribution of 32% and 60% of the total ecosystem carbon residing in the soil reservoir for tropical and boreal forests, respectively (Pan et al., 2011). This vast difference in soil carbon pools between ecosystems reflects differences in the intensity of processes that lead to accumulation and stabilization of carbon in soils. If climate change stimulates root growth (Norby and Zak, 2011), then the amount of carbon allocated to symbionts like mycorrhizae will also increase, leading to increased amounts of stabilized soil carbon via components such as chitin and glomalin (Treseder and Allen, 2000). It is estimated that 41% of belowground NPP is allocated to ectomycorrhizal symbionts, and more significantly, it is this symbiotic relationship that is believed to be the main process by which boreal and temperate forest soils have accumulated carbon (Hobbie, 2006). Carbon accumulation and storage in soils, especially deeper in soil profiles, is a slow process that progresses over centuries and millennia, although certain management practices and disturbances such as land use change can cause soil carbon in much shorter periods (Jandl et al., 2007). In forest systems most carbon is sequestered in the form of plant biomass, with only a small portion (around 33 g C m2 y1) transferred to the soil (Post and Kwon, 2000); of this, only 0.2e1.2 g C m2 y1 represent stabilized soil carbon (Schlesinger et al., 2000). In boreal ecosystems the change from forest to agriculture can reduce the soil carbon reservoir by 8% in one decade, with even higher losses possible in soils under permafrost (Gruenzweig et al., 2015). Compared to boreal ecosystems, tropical forests store less soil carbon, and their carbon stocks in peat and secondary vegetation are strongly linked to global cyclic events like El Ninõ; for example, in the drought of 1997, burning of peat and vegetation in Indonesia was equivalent to 13%e40% of average global fossil fuel emissions (Page et al., 2002). The amount of carbon sequestered in forest soils is influenced by tree species. For soils with similar mineralogy, the rate of production of detritus and its quality determine how much carbon is stored in soil. Fig. 7.2 shows accumulation of new soil carbon with time upon reforestation of cultivated

194


New soil C (Mg/ha)

20 y = 0.039x2 + 0.063x R² = 0.829

0-5 cm

15 y = 0.024x + 0.004x R² = 0.575

5-10 cm

10 5 0 0

5

10 15 20 Time aŌer reforestaƟon (y)

25

30

FIGURE 7.2 Soil carbon accumulation following reforestation with three pine species in andisols of Central Mexico. Soils were previously cultivated under corn, and new soil carbon was estimated from changes in 13C composition using a standard isotopic mixing model and a chronosequence of reforestations (Unpublished data).

Table 7.1 New Soil Carbon Accumulation (Mg ha1 y1) for Individual Pine Species Used to Reforest Andisols Previously Cultivated Under Corn Soil Depth

P. devoniana

P. patula

P. greggii

0e5 cm

0.561

0.444

0.545

5e10 cm

0.115

0.174

0.301

Total from 0 to 10

0.676

0.618

0.846

andisols with three pine species in Central Mexico. The specific rates of accumulation under each species are given in Table 7.1. Andisols show high rates of soil formation (Osman, 2013) and are capable of storing relatively large amounts of carbon because they are rich in noncrystalline minerals. However, not all carbon stored in recently reforested soils is stabilized carbon. In evaluating long-term carbon sequestration, the stabilized fraction of soil carbon is most important; a frame of reference for carbon that may be considered “stabilized” has been described for preserved grasslands as that which remains for 97 years (Torn et al., 2002). Soil carbon dynamics and stabilization are influenced not only by tree species but also by understory species, on the condition that an N-fixing species is present; soils lose carbon faster when understory is removed or when the organic inputs to the soil come from a single species (Winsome et al., 2007). Thus as climate change influences understory diversity, some changes in the dynamics of soil carbon and nitrogen are expected. In addition to characteristics of soils and tree species, the capacity of forest soils to accumulate and store carbon, and therefore the overall carbon cycle of the system, is influenced by human pressure on forest goods. For example, in semiarid systems in northeast Brazil, firewood harvest is higher than NPP.

The Capacity of Forest Soils to Provide Ecosystem Services

195

Simulations suggest that if harvest of firewood stops, the recovery time for this system to return to its initial soil carbon content is 50 years under current conditions. In contrast, if climate change brings lower rainfall and higher temperatures, continued harvest of firewood will reduce soil carbon by 13% in this century (Althoff et al., 2016). The effect of climate change on soil carbon sequestration and storage in forest ecosystems will depend on the concurrent effects on NPP, on soil properties and processes (e.g., the balance between stabilization and mineralization of carbon), and on changes in external influences (e.g., tree harvest).

THE CAPACITY OF FOREST SOILS TO PROVIDE ECOSYSTEM SERVICES The soil is the supporting component for all the ecosystem services that forests provide to humans. All forest ecosystems worldwide, and their ability to provide environmental services, have been affected by human influence (Pan et al., 2013). Changes in climate can further influence the degree to which forests provide these environmental services. For example, it has been suggested that boreal forests, some of the largest forests on Earth, will in time lose their ability to sequester carbon as an environmental service, becoming instead a source of C (Gauthier et al., 2015). Although the combination of increased CO2 and temperature is expected to promote tree growth in these forests, the effect may not be sustained over time; furthermore, at some point there will be limitations in nutrient and water availability (Stinziano and Way, 2014). The relatively large pools of C as litter and soil in boreal forests will amplify any effect of climate change on increasing release of C from these systems (Kurz et al., 2013). The greater variability in temperature and precipitation expected as a result of climate change will affect biotic processes in forest soils such as nutrient and carbon mineralization, immobilization, and respiration. These processes, in combination with abiotic phenomena related to the adsorption capacity of the mineral soil, will have direct effects on the provision of environmental services (Campbell et al., 2009). Environmental services related to cycling and release of harmful elements are important benefits of forests in certain areas. For example, climate change is expected to impact forests in the northeastern United States by favoring the transition from conifers to deciduous species, which in turn can affect the mercury (Hg) cycle in this region. The forest floor under coniferous species (Abies balsamea Mill.; Picea rubens Sarg.; Tsuga canadensis L.) accumulates 40% more Hg compared to the forest floor under deciduous species (Acer saccharum; Marsh.; Acer rubrum L.; Acer pensylvanicum Marsh; Betula papyrifera; Marsh.; Betula alleghaniensis Britt); however, fluxes of Hg from canopy to soil are higher in deciduous forests because annual production of litter is eight times that of coniferous forests (Richardson and Friedland, 2015). Some authors, in contrast, have reported greater accumulation of Hg in the soil of

196


deciduous forests compared with coniferous forests, and higher net inputs of Hg in coniferous forests with low canopy cover due to higher inputs of Hg in throughfall (Blackwell et al., 2014). Although forest floor and soil represent different stages in the environmental movement of Hg, these contrasting findings indicate that the potential effects of climate change on the Hg cycle, and likely other elemental cycles, are still uncertain. Furthermore, other factors control the movement of Hg in the environment, including sorption as influenced by soil texture and organic matter (Richardson and Friedland, 2015), and export of Hg to groundwater as influenced by organic matter composition and solubility (Blackwell et al., 2014). Some environmental services, such as those related to water quality, may improve with climate change. It has been suggested that higher temperatures and decreased winter rainfall in forests of Picea abies Karst. and Fagus sylvatica L. will lead to a reduction of nitrate in water. Certain management practices can enhance this benefit; for example, although clear cutting and shelterwood practices initially cause the concentration of nitrate in streams to peak, annual net outputs of nitrate are lower due to a higher number of young trees in the forest structure that can efficiently take up any excess nitrate (Thomas et al., 2016). Higher temperatures would stimulate mineralization of soil nitrogen leading to the release of more inorganic N, but this response would be constrained by limitations in soil water content. It is notable that an increase of 5 C in the soil forests ecosystems is estimated to lead to additional inorganic N production of up to 80 kg ha1 y1, equal to eight times the atmospheric inputs of nitrogen observed in pristine forests (Campbell et al., 2009). Although temperature and moisture are important factors that determine mineralization of soil nitrogen, the quality and total amount of substrate are important as well (Campbell et al., 2009); these too can be affected by climate change.

SOIL PROCESSES IN RELATION TO SOIL TEXTURE Soil texture is an important soil characteristic that can modulate the effects of climate change via its influence on components of the carbon cycle in forests, including tree growth response and soil organic matter retention. Fine particles have higher specific surface area and are more reactive than coarse particles, therefore clay-textured soils generally store higher amounts of carbon than sandy soils (Sulman et al., 2014). The combination of readily decomposable litter and high clay content in forests leads to the greatest amounts of stabilized carbon in soil (Sulman et al., 2014). However, in addition to soil texture, mineralogy also affects the amount of carbon ultimately stabilized in soil, particularly the amount of noncrystalline minerals; indeed, it is this property that often explains differences in the amount of carbon stored in soil profiles (e.g., Torn et al., 1997; Rasmussen et al., 2006). Microbial activity in soil, which is also related to carbon stabilization and storage, is influenced by the effect of texture on water holding capacity and aeration. For example, under similar conditions of temperature and moisture, respiration can be higher in fine-textured soils

Soil Processes in Relation to Soil Texture 1.2

2.5 2.0

Clay

1.5

Sandy Loam

mg CO2-C /g soil

Loam mg CO2-C /g soil

197

1.0 0.5

Loam

1.0

Clay

0.8

Sandy Loam

0.6 0.4 0.2 0.0

0.0 0

20

40

60 Days

80

100

0

20

40

60

80

100

Days

FIGURE 7.3 CO2 mineralization for the surface soil at two soil depths, 0e15 (left) and 15e30 (right) for three forests soils (Haploxeralfs, Palexerults, and Dystrochrepts) with different soil texture (Unpublished data). The four lines for each soil texture represent soils with different organic carbon content. For the silvicultural history and mineralogy background of these soils see Gomez et al. (2002); Powers et al. (2005) and Ramussen et al. (2006).

(Jana et al., 2010), due to higher amounts of substrate per unit mass of soil, as well as higher water retention. In addition to the overall capacity to store water, soil texture also influences NPP through its control of water availability, according to precipitation and the final balance of air and water in soil (Gomez et al., 2002: Delgado-Caballero et al., 2009). To demonstrate the effect of soil texture on microbial respiration, Fig. 7.3 presents data for CO2 evolution measured in three soils representative of Northern California forests at a constant temperature (22 C) and relative water content (50% of water holding capacity). The finer-textured soils (clay and loam) mineralized more carbon than the sandy loam. After 102 days of incubation, variation in carbon mineralization for all soils ranged from 0.5 to 1.9 mg CO2-C g1 for the top soil and from 0.4 to 1.0 for the second depth. The amount of total organic carbon in these soils follows the order loam > clay > sandy loam, which can be explained in part by parent materials, respectively, of andesite, basalt, and granite (Powers et al., 2005; Rasmussen et al., 2006). As expected, there is a correlation between soil organic carbon content and the amount of CO2 produced. The effect of texture and mineralogy, however, is evident if the amount of CO2 involved is expressed as a percent of total soil carbon (data not shown), which followed the order clay > sandy loam > loam. The influence of texture and mineralogy on retention and loss of soil carbon is therefore important to consider when evaluating the impact that climate change might have on a given forest ecosystem, and when setting the parameters of models designed to predict these impacts. An example of the role of soil texture in modulating the effects of climate change has been observed among Quercus suber forests in southern Spain (Ibanez et al., 2014), the survival of which is related to soil moisture, itself controlled by texture. Survival of adult trees increased in locations with higher winter rainfall and sandy surface soils, as infiltration is higher, permitting greater water storage in deeper soils. In contrast, sites with clayey surface soils have lower infiltration and were not able to store as much water. The water soil

198


potential can also be important in this relationship with more negative potentials in clay soils due to the matrix potential effect. Another indirect effect of texture is its influence on the amount of solar energy retained by the surface soil. In boreal forests, sandy and well-drained soils show higher surface temperatures than finer soils with poor drainage, which results in higher rates of organic matter decomposition in Pinus banksiana compared to Picea marina stands in northern Canada (Preston et al., 2006). The differences in surface soil temperature between these stands is 2e3 C, suggesting that climate change may exert different effects on different forests, with consequent profound impacts on the carbon cycle (Preston et al., 2006). Short-term incubations have demonstrated the importance of factors such as temperature and texture in influencing the amount of soil carbon readily lost through mineralization, but long-term incubations provide additional insight into the factors that control loss of stabilized carbon, such as availability of nitrogen (Tian et al., 2016). Furthermore, the dynamics of decomposition can be described in greater detail; for example, the sand fraction of soil has been found to be related to the fraction of soil carbon with an intermediate rate of decomposition, since compared to finer particles, sands have lower sorption capacity, which leaves more carbon susceptible to decomposition (Tian et al., 2016). Similarly, the relative proportion of soil carbon pools differing in decomposition rate has been correlated to clay content, again highlighting the protective effect of clay (Chiti et al., 2012; Singh et al., 2011). The importance of soil texture on forest productivity and carbon and nitrogen cycling is evident in moist tropical and arid environments as well. A positive correlation has been documented in the Amazon forests between clay content and NPP (Laurance et al., 1999), a relationship that can be explained by greater amounts of organic matter and nitrogen in clay soils. In desert soils, where rainfall dynamics are much different, there is an interesting effect of texture on soil respiration: losses of C via heterotrophic respiration are higher in fine-textured soils, while the response of autotrophic respiration to pulses of rain are higher in coarse soils (Cable et al., 2008). The effects of climate change in arid environments, therefore, will not only depend on the effect of soil texture but also on the temporal distribution of precipitation. Soil texture is therefore a useful parameter to consider when predicting the effects of climate change on forest ecosystems, not only because of its influence on important variables such as water content but also because of the correlations between texture and soil carbon dynamics and vegetation distribution (Haxeltine et al., 1996; Schimel et al., 1994)

MICROBIAL PROCESSES IN FOREST SOILS Global soil carbon stocks are estimated at 1500 Pg, of which about 4% (60 Pg) are respired annually (Schlesinger and Bernhardt, 2013). Carbon inputs to forest soils are governed by NPP of vegetation, the amount of litterfall, and microbial activity (Schlesinger and Bernhardt, 2013), while outputs or losses occur

Microbial Processes in Forest Soils

199

100

Loam Clay

80

Sandy loam 60 40 20 0 0

50

100 Days

150

N Mineralization (mg kg soil)

N Mineralization (mg kg soil)

via soil respiration, movement of dissolved organic carbon, and disturbances such as deforestation, fire, and erosion. As discussed earlier, the soil as a physical system exhibits specific characteristics that control biological processes involved in accumulation and loss of soil carbon, including texture, structure, porosity, bulk density, and water holding capacity. In general, under similar water content and temperature, soils with larger pools of carbon respire more CO2 and show concurrently higher rates of nitrogen mineralization. For example, soil respiration rates for Betula forests in Scotland were about six times higher than in Pinus forests, which have about 10 times less soil carbon; expressed relative to soil carbon content, respiration is 3.3% and 0.34%, respectively, in Betula and Pinus forests (Nazaries et al., 2015). It has been estimated that of the total respiration in forest soils, from 18% to 39% is due to hyphae and from 16% to 29% can be attributed to root systems (Andrew et al., 2014). Regarding the effect of temperature, laboratory incubations and study of soil microbial communities in a variety of different soils, including forests and grasslands, have shown that although the composition of the microbial community changed upon an increase in temperature from 7 to 10 C, respiration rate nevertheless increased under all land uses (e.g., Nazaries et al., 2015). Similar to carbon mineralization, nitrogen mineralization is dependent on soil microbial activity, organic matter content, and substrate quality and availability. Nitrogen mineralization is shown in Fig. 7.4 for the same three soils described in Fig. 7.3, which range in carbon content from 1.3% to 7.8%, and a similar C:N ratio of 20 in all soils. Nitrogen mineralization mimics carbon mineralization, with higher mineralization in fine-textured soils compared to coarse soils. In addition to carbon content, this observation is due to the fact that finertextured soils have a more complex microstructure to host microbial communities, allowing mineralization to proceed more rapidly. As is the case for soil carbon dynamics, such differences in release of nitrogen (and other nutrients) as a function of soil properties is important to consider when constructing predictive models. It has been shown that indices of aridity together with substrate availability can explain observed heterotrophic respiration (Fang et al., 2014). Climate change will affect not only the overall activity but also the composition of microbial populations in forest soils. As soil moisture decreases,

40 Loam Clay

30

Sandy Loam 20 10 0 0

50

100

150

Days

FIGURE 7.4 Nitrogen mineralization for the surface soil (0e15 cm) and subsoil (15e30) for the same three forest soils described in Fig. 7.3 (unpublished data). The four lines for each soil texture represent soils with different organic carbon content.

200


microbial activity, root exudation, and mass transport and diffusion of nutrients also decrease (Gessler et al., 2007; Gschwendtner et al., 2014). Rewetting can lead to pulses in activity; the population of denitrifiers, for example, has been observed to increase after soil is rewetted (Gschwendtner et al., 2014). If climate change leads to more pronounced and more frequent recurrence of drought and wetting, this will likely result in more release of NO and N2O to the atmosphere. Under intensive management, the practice of harvesting whole trees can lead to increased N2O emissions if soil water content increases and the soil environment becomes more reducing (McDaniel et al., 2014). Overall, of the two main factors that control soil microbial activitydmoisture and temperaturedsoil moisture appears to be more important (Gessler et al., 2007; Gschwendtner et al., 2014; McDaniel et al., 2014). However, these two variables are related, and after a certain degree of warming, the soil loses moisture and microbial activity declines. In forests that receive heavy snowfall, higher winter temperatures may increase the frequency of freezeethaw cycles, in turn promoting soil nitrogen mineralization and release of dissolved organic carbon. If relatively little snow accumulates on a forest soil, the number of freezeethaw cycles may decrease, causing higher soil temperatures, variations in microbial populations, and increased losses of carbon as well (Urakawa et al., 2014). As an example of how soil properties can dictate microbial community structure, results from a phospholipid fatty acid profile of the same soils described in Figs. 7.3 and 7.4, as well as a reference soil to compare to natural forest, are presented in Fig. 7.5. Characterization of the fatty acid profile was based on studies of the composition of fatty acids in ectomycorrhizae (Karli nski et al., 2007), studies carried out in forest soils (Arunkumar et al., 2015; Karli nski et al., 2007; Moore-Kucera and Dick, 2008; Overby et al., 2015), and the considerations regarding interpretation as explained by Frostegard et al. (2011). The data indicate a higher relative presence of gram-negative bacteria in the sandy loam compared to the other textures, while the fungal population and the ratio of fungi to bacteria were higher

0.5

20 15

a

ab

b

10

a

b

ab

ab

ab

5

Fungi / Bacterial

% of total PFLA

25 0.4

a b

ab

ab

0.3 0.2 0.1 0

0 Loam Gram (-)

Clay Gram (+)

Sand Loam AcƟnomycetes

Reference Soil Fungi

Loam

Clay

Sand Loam Reference Soil

FIGURE 7.5 Microbial community composition described by phospholipid fatty acid profiles (left) and fungi to bacteria ratio (right) in the surface soil (0e15 cm) of the three forest soils described in Figs. 7.3 and 7.4 (unpublished data). Different letters indicate a significant difference (P < .05) across soil type according to the Duncan test.

References

201

in the clay soil. As mentioned earlier, the physical composition a soil (e.g., the clay content) appears to be important in defining the microbial community structure. The reference soil, undisturbed soil under coniferous mixed forest, did not differ strongly from the other soils, which are managed plantations. It has been previously reported that microbial populations in forest plantations show a microbial community structure similar to that of natural forests after several years to decades (Moore-Kucera and Dick, 2008; Okland et al., 2004).

CONCLUSIONS As seen from the present discussion, the effects of climate change on forests and the processes that occur in forest soils are complex. Although much research has been carried out on different facets of these interrelationships, experimental results do not always confirm what is observed naturally in systems different from those studied or over longer time periods, and certain effects remain difficult to predict due to the interplay of multiple factors. In spite of this, some recurring findings are evident, such as the conclusion that the extent of the effect of climate change in a given system will depend on the extent to which soil water availability is altered. Temperature and water availability are also strongly related in this regard, with the effect of increasing temperature reaching a critical threshold for tree growth once evapotranspiration begins to limit soil water availability. Before this stage is reached, soil biological processes and tree growth can change favorably. Additional consideration must be given to the changes imposed by disturbances such as alterations in the nitrogen cycle, air pollution, and shifts in management; the net effect of two or more factors may be synergistic in causing changes in the system as a whole. A commonly reported system-wide change is a shift in species composition, which in turn changes litter quality and the way in which nutrients are cycled. Environmental services of forest ecosystems, such as maintenance of water quality, are also expected to be modified by climate change, either positively or negatively depending on the nature of the system. Perhaps the foremost and most widely recognized indicator of changes in forest ecosystems is net productivity, which integrates the overall effect of processes ranging from carbon fixation and allocation to nutrient cycling between canopy and soil. All of these processes are interconnected, underscoring the value of insightful observations of trees and soil processes, and especially the connections between them, as part of an improved understanding of how forest ecosystems will respond to climate change.

REFERENCES Althoff, T.D., Menezes, R.S.C., de Carvalho, A.L., de Siqueira Pinto, A., Santiago, G.A.C.F., Ometto, J.P.H.B., von Randow, C., Sampaio, E.V.d.S.B., 2016. Climate change impacts on the sustainability of the firewood harvest and vegetation and soil carbon stocks in a tropical dry forest in Santa Teresinha Municipality, Northeast Brazil. Forest Ecology and Management 360, 367e375.

202


Andrew, C.J., Van Diepen, L.T.A., Miller, R.M., Lilleskov, E.A., 2014. Aspen-associated mycorrhizal fungal production and respiration as a function of changing CO2, O3 and climatic variables. Fungal Ecology 10, 70e80. Arunkumar, R.D.S., Patra, A.K., Sahu, S.K., Sopan, B., Bhople, M.A.K., Nimje, A., 2015. Impact of canopy and seasonal dynamics on oak and pine soil microbial community composition and other biological indicators in Garhwal Himalayas. Ecology Environment and Conservation 21, S213eS226. Banin, L., Lewis, S.L., Lopez-Gonzalez, G., Baker, T.R., Quesada, C.A., Chao, K.J., Burslem, D.F., Nilus, R., Abu Salim, K., Keeling, H.C., 2014. Tropical forest wood production: a cross-continental comparison. Journal of Ecology 102 (4), 1025e1037. Berner, L.T., Law, B.E., Hudiburg, T.W., 2017. Water availability limits tree productivity, carbon stocks, and carbon residence time in mature forests across the western US. Biogeosciences 14 (2), 365e378. Blackwell, B.D., Driscoll, C.T., Maxwell, J.A., Holsen, T.M., 2014. Changing climate alters inputs and pathways of mercury deposition to forested ecosystems. Biogeochemistry (Dordrecht) 119 (1e3), 215e228. Breymeyer, A., 2003. Rate of change of pine litter fall and decay in response to changing climate along N-S transect. Polish Journal of Ecology 51 (4), 553e555. Cable, J.M., Ogle, K., Williams, D.G., Weltzin, J.F., Huxman, T.E., 2008. Soil texture drives responses of soil respiration to precipitation pulses in the Sonoran Desert: implications for climate change. Ecosystems 11 (6), 961e979. Camarero, J.J., Gazol, A., Galvań, J.D., Sanguësa-Barreda, G., Gutie´rrez, E., 2015. Disparate effects of global-change drivers on mountain conifer forests: warming-induced growth enhancement in young trees vs. CO2 fertilization in old trees from wet sites. Global Change Biology 21 (2), 738e749. Campbell, J.L., Rustad, L.E., Boyer, E.W., Christopher, S.F., Driscoll, C.T., Fernandez, I.J., Groffman, P.M., Houle, D., Kiekbusch, J., Magill, A.H., 2009. Consequences of climate change for biogeochemical cycling in forests of northeastern North America. Canadian Journal of Forest Research 39 (2), 264e284. Chertov, O., Bhatti, J.S., Komarov, A., 2010. Impact of temperature increase and precipitation alteration at climate change on forest productivity and soil carbon in boreal forest ecosystems in Canada and Russia: simulation approach with the EFIMOD model. In: Simard, S.W., Austin, M.E. (Eds.), Climate Change and Variability, pp. 303e326. Chiti, T., Dıáz-Pine´s, E., Rubio, A., 2012. Soil organic carbon stocks of conifers, broadleaf and evergreen broadleaf forests of Spain. Biology and Fertility of Soils 48 (7), 817e826. Davi, H., Dufrene, E., Francois, C., Le Maire, G., Loustau, D., Bosc, A., Rambal, S., Granier, A., Moors, E., 2006. Sensitivity of water and carbon fluxes to climate changes from 1960 to 2100 in European forest ecosystems. Agricultural and Forest Meteorology 141 (1), 35e56. Davies, J., Tipping, E., Whitmore, A., 2016. 150 years of macronutrient change in unfertilized UK ecosystems: observations vs simulations. Science of the Total Environment 572, 1485e1495. Delgado-Caballero, C.E., Go´mez-Guerrero, A., Valdez-Lazalde, J.R., los Santos-Posadas, D., Fierros-Gonza´lez, A.M., Horwath, W.R., 2009. Site index and soil properties in young plantations of Eucalyptus grandis and E. urophylla in Southeastern Mexico. Agrociencia 43 (1), 61e72. De Marco, A., Proietti, C., Cionni, I., Fischer, R., Screpanti, A., Vitale, M., 2014. Future impacts of nitrogen deposition and climate change scenarios on forest crown defoliation. Environmental Pollution 194, 171e180. Drake, J.E., Macdonald, C.A., Tjoelker, M.G., Crous, K.Y., Gimeno, T.E., Singh, B.K., Reich, P.B., Anderson, I.C., Ellsworth, D.S., 2016. Short-term carbon cycling responses of a mature eucalypt woodland to gradual stepwise enrichment of atmospheric CO2 concentration. Global Change Biology 22 (1), 380e390.

References

203

Duursma, R.A., Gimeno, T.E., Boer, M.M., Crous, K.Y., Tjoelker, M.G., Ellsworth, D.S., 2016. Canopy leaf area of a mature evergreen Eucalyptus woodland does not respond to elevated atmospheric [CO2] but tracks water availability. Global Change Biology 22, 1666e1676. Fang, H., Cheng, S., Wang, Y., Yu, G., Xu, M., Dang, X., Li, L., Wang, L., 2014. Changes in soil heterotrophic respiration, carbon availability, and microbial function in seven forests along a climate gradient. Ecological Research 29 (6), 1077e1086. Feng, Z., Ru¨tting, T., Pleijel, H., Wallin, G., Reich, P.B., Kammann, C.I., Newton, P.C., Kobayashi, K., Luo, Y., Uddling, J., 2015. Constraints to nitrogen acquisition of terrestrial plants under elevated CO2. Global Change Biology 21 (8), 3152e3168. Fenn, M.E., Huntington, T., McLaughlin, S., Eagar, C., Gomez, A., Cook, R., 2006. Status of soil acidification in North America. Journal of Forest Science (Prague) 52, 3e13. Friedlingstein, P., Houghton, R.A., Marland, G., Hackler, J., Boden, T.A., Conway, T.J., Canadell, J.G., Raupach, M.R., Ciais, P., Le Quere, C., 2010. Update on CO2 emissions. Nature Geoscience 3 (12), 811e812. ˚ ., Tunlid, A., Ba˚a˚th, E., 2011. Use and misuse of PLFA measurements in soils. Frostega˚rd, A Soil Biology and Biochemistry 43 (8), 1621e1625. Gauthier, S., Bernier, P., Kuuluvainen, T., Shvidenko, A.Z., Schepaschenko, D.G., 2015. Boreal forest health and global change. Science 349 (6250), 819e822. Ge, Z.-M., Kellomaki, S., Peltola, H., Zhou, X., Vaisanen, H., Strandman, H., 2013. Impacts of climate change on primary production and carbon sequestration of boreal Norway spruce forests: Finland as a model. Climatic Change 118 (2), 259e273. Gessler, A., Keitel, C., Kreuzwieser, J., Matyssek, R., Seiler, W., Rennenberg, H., 2007. Potential risks for European beech (Fagus sylvatica L.) in a changing climate. Trees (Berlin) 21 (1), 1e11. Gomez, A., Powers, R., Singer, M., Horwath, W., 2002. Soil compaction effects on growth of young ponderosa pine following litter removal in California’s Sierra Nevada. Soil Science Society of America Journal 66 (4), 1334e1343. Gruenzweig, J.M., Valentine, D.W., Chapin III, F.S., 2015. Successional changes in carbon stocks after logging and deforestation for agriculture in interior Alaska: implications for boreal climate feedbacks. Ecosystems 18 (1), 132e145. Grulke, N., Tausz, M., 2014. The future of trees in a changing climate: synopsis. In: Tausz, M., Grulke, N. (Eds.), Trees in a Changing Environment: Ecophysiology, Adaptation, and Future Survival. Springer Netherlands, Dordrecht, pp. 265e279. Gschwendtner, S., Tejedor, J., Bimueller, C., Dannenmann, M., Knabner, I.K., Schloter, M., 2014. Climate change induces shifts in abundance and activity pattern of bacteria and archaea catalyzing major transformation steps in nitrogen turnover in a soil from a mid-European beech forest. PLoS One 9 (12), e114278. Gueńon, R., Day, T.A., Velazco-Ayuso, S., Gros, R., 2017. Mixing of Aleppo pine and Holm oak litter increases biochemical diversity and alleviates N limitations of microbial activity. Soil Biology and Biochemistry 105, 216e226. Handa, I.T., Korner, C., Hattenschwiler, S., 2006. Conifer stem growth at the altitudinal treeline in response to four years of CO2 enrichment. Global Change Biology 12 (12), 2417e2430. Haxeltine, A., Prentice, I.C., Creswell, I.D., 1996. A coupled carbon and water flux model to predict vegetation structure. Journal of Vegetation Science 7 (5), 651e666. Hobbie, E.A., 2006. Carbon allocation to ectomycorrhizal fungi correlates with belowground allocation in culture studies. Ecology 87 (3), 563e569. Ibanez, B., Ibanez, I., Gomez-Aparicio, L., Ruiz-Benito, P., Garcia, L.V., Maranon, T., 2014. Contrasting effects of climate change along life stages of a dominant tree species: the importance of soil-climate interactions. Diversity and Distributions 20 (8), 872e883. Jana, B.K., Biswas, S., Sonkar, S., Majumder, M., Roy, P., Mazumdar, A., 2010. Estimation of soil carbon stock and soil respiration rate of recreational and natural forests in India. In: Impact of Climate Change on Natural Resource Management.

204


Jandl, R., Lindner, M., Vesterdal, L., Bauwens, B., Baritz, R., Hagedorn, F., Johnson, D.W., Minkkinen, K., Byrne, K.A., 2007. How strongly can forest management influence soil carbon sequestration? Geoderma 137 (3), 253e268. Jones, A.G., Scullion, J., Ostle, N., Levy, P.E., Gwynn-Jones, D., 2014. Completing the FACE of elevated CO2 research. Environment International 73, 252e258. Karlinski, L., Ravnskov, S., Kieliszewska-Rokicka, B., Larsen, J., 2007. Fatty acid composition of various ectomycorrhizal fungi and ectomycorrhizas of Norway spruce. Soil Biology and Biochemistry 39 (4), 854e866. Kirschbaum, M.U.F., Watt, M.S., Tait, A., Ausseil, A.-G.E., 2012. Future wood productivity of Pinus radiata in New Zealand under expected climatic changes. Global Change Biology 18 (4), 1342e1356. Krishnaswamy, J., John, R., Joseph, S., 2014. Consistent response of vegetation dynamics to recent climate change in tropical mountain regions. Global Change Biology 20 (1), 203e215. Kurz, W.A., Shaw, C., Boisvenue, C., Stinson, G., Metsaranta, J., Leckie, D., Dyk, A., Smyth, C., Neilson, E., 2013. Carbon in Canada’s boreal forestd a synthesis 1. Environmental Reviews 21, 260e292. Laurance, W.F., Fearnside, P.M., Laurance, S.G., Delamonica, P., Lovejoy, T.E., Rankin-de Merona, J.M., Chambers, J.Q., Gascon, C., 1999. Relationship between soils and Amazon forest biomass: a landscape-scale study. Forest Ecology and Management 118 (1), 127e138. Lu, X., Jiang, H., Liu, J., Zhang, X., Jin, J., Zhu, Q., Zhang, Z., Peng, C., 2016. Simulated effects of nitrogen saturation on the global carbon budget using the IBIS model. Scientific Reports 6. Makipaa, R., Linkosalo, T., Komarov, A., Makela, A., 2015. Mitigation of climate change with biomass harvesting in Norway spruce stands: are harvesting practices carbon neutral? Canadian Journal of Forest Research 45 (2), 217e225. McDaniel, M.D., Kaye, J.P., Kaye, M.W., 2014. Do "hot moments" become hotter under climate change? Soil nitrogen dynamics from a climate manipulation experiment in a post-harvest forest. Biogeochemistry (Dordrecht) 121 (2), 339e354. McTiernan, K.B., Couteaux, M.-M., Berg, B., Berg, M.P., de Anta, R.C., Gallardo, A., Kratz, W., Piussi, P., Remacle, J., De Santo, A.V., 2003. Changes in chemical composition of Pinus sylvestris needle litter during decomposition along a European coniferous forest climatic transect. Soil Biology & Biochemistry 35 (6), 801e812. Moore-Kucera, J., Dick, R.P., 2008. PLFA profiling of microbial community structure and seasonal shifts in soils of a Douglas-fir chronosequence. Microbial Ecology 55 (3), 500e511. Munir, T.M., Xu, B., Perkins, M., Strack, M., 2014. Responses of carbon dioxide flux and plant biomass to water table drawdown in a treed peatland in northern Alberta: a climate change perspective. Biogeosciences 11 (3), 807e820. Nazaries, L., Tottey, W., Robinson, L., Khachane, A., Abu Al-Soud, W., Sorensen, S., Singh, B.K., 2015. Shifts in the microbial community structure explain the response of soil respiration to land-use change but not to climate warming. Soil Biology & Biochemistry 89, 123e134. Noormets, A., Epron, D., Domec, J.-C., McNulty, S., Fox, T., Sun, G., King, J.S., 2015. Effects of forest management on productivity and carbon sequestration: a review and hypothesis. Forest Ecology and Management 355, 124e140. Norby, R.J., Zak, D.R., 2011. Ecological lessons from free-air CO2 enrichment (FACE) experiments. Annual Review of Ecology, Evolution, and Systematics 42 (1), 181. Ogaya, R., Llusia, J., Barbeta, A., Asensio, D., Liu, D., Arturo Alessio, G., Penuelas, J., 2014. Foliar CO2 in a holm oak forest subjected to 15 years of climate change simulation. Plant Science (Oxford) 226 (Sp. Iss. SI), 101e107. Okland, T., Bakkestuen, V., Okland, R.H., Eilertsen, O., 2004. Changes in forest understory vegetation in Norway related to long-term soil acidification and climatic change. Journal of Vegetation Science 15 (4), 437e448.

References

205

Oren, R., Ellsworth, D.S., Johnsen, K.H., Phillips, N., Ewers, B.E., Maier, C., Scha¨fer, K.V., McCarthy, H., Hendrey, G., McNulty, S.G., 2001. Soil fertility limits carbon sequestration by forest ecosystems in a CO2-enriched atmosphere. Nature 411 (6836), 469e472. Osman, K.T., 2013. Soils: Principles, Properties and Management. Springer Science & Business Media, Dordrecht, 271 p. Overby, S.T., Owen, S.M., Hart, S.C., Neary, D.G., Johnson, N.C., 2015. Soil microbial community resilience with tree thinning in a 40-year-old experimental ponderosa pine forest. Applied Soil Ecology 93, 1e10. Page, S.E., Siegert, F., Rieley, J.O., Boehm, H.-D.V., Jaya, A., Limin, S., 2002. The amount of carbon released from peat and forest fires in Indonesia during 1997. Nature 420 (6911), 61e65. Pan, Y., Birdsey, R.A., Fang, J., Houghton, R., Kauppi, P.E., Kurz, W.A., Phillips, O.L., Shvidenko, A., Lewis, S.L., Canadell, J.G., 2011. A large and persistent carbon sink in the world’s forests. Science 333 (6045), 988e993. Pan, Y., Birdsey, R.A., Phillips, O.L., Jackson, R.B., 2013. The structure, distribution, and biomass of the world’s forests. Annual Review of Ecology, Evolution, and Systematics 44, 593e622. Petrie, M.D., Pockman, W.T., Pangle, R.E., Limousin, J.M., Plaut, J.A., McDowell, N.G., 2015. Winter climate change promotes an altered spring growing season in pinon pinejuniper woodlands. Agricultural and Forest Meteorology 214, 357e368. Post, W.M., Kwon, K.C., 2000. Soil carbon sequestration and land-use change: processes and potential. Global Change Biology 6 (3), 317e327. Poulter, B., Pederson, N., Liu, H., Zhu, Z., D’Arrigo, R., Ciais, P., Davi, N., Frank, D., Leland, C., Myneni, R., Piao, S., Wang, T., 2013. Recent trends in Inner Asian forest dynamics to temperature and precipitation indicate high sensitivity to climate change. Agricultural and Forest Meteorology 178, 31e45. Powers, R.F., Scott, D.A., Sanchez, F.G., Voldseth, R.A., Page-Dumroese, D., Elioff, J.D., Stone, D.M., 2005. The North American long-term soil productivity experiment: findings from the first decade of research. Forest Ecology and Management 220 (1), 31e50. Preston, C.M., Bhatti, J.S., Flanagan, L.B., Norris, C., 2006. Stocks, chemistry, and sensitivity to climate change of dead organic matter along the Canadian boreal forest transect case study. Climatic Change 74 (1e3), 223e251. Priha, O., Grayston, S.J., Hiukka, R., Pennanen, T., Smolander, A., 2001. Microbial community structure and characteristics of the organic matter in soils under Pinus sylvestris, Picea abies and Betula pendula at two forest sites. Biology and Fertility of Soils 33 (1), 17e24. Rasmussen, C., Southard, R.J., Horwath, W.R., 2006. Mineral control of organic carbon mineralization in a range of temperate conifer forest soils. Global Change Biology 12 (5), 834e847. Richardson, J.B., Friedland, A.J., 2015. Mercury in coniferous and deciduous upland forests in northern New England, USA: implications of climate change. Biogeosciences 12 (22), 6737e6749. Santonja, M., Fernandez, C., Gauquelin, T., Baldy, V., 2015. Climate change effects on litter decomposition: intensive drought leads to a strong decrease of litter mixture interactions. Plant and Soil 393 (1e2), 69e82. Schimel, D.S., Braswell, B., Holland, E.A., McKeown, R., Ojima, D., Painter, T.H., Parton, W.J., Townsend, A.R., 1994. Climatic, edaphic, and biotic controls over storage and turnover of carbon in soils. Global Biogeochemical Cycles 8 (3), 279e293. Schindlbacher, A., Wunderlich, S., Borken, W., Kitzler, B., Zechmeister-Boltenstern, S., Jandl, R., 2012. Soil respiration under climate change: prolonged summer drought offsets soil warming effects. Global Change Biology 18 (7), 2270e2279. Schlesinger, W.H., Bernhardt, E.S., 2013. Biogeochemistry: An Analysis of Global Change, third ed. Academic Press, San Diego, California. Schlesinger, W.H., Winkler, P.J., Megonigal, J.P., 2000. Ecological studies. In: Wigley, T., Schimel, D. (Eds.), The Carbon Cycle. Cambridge University Press, Cambridge, pp. 96e101.

206


Sicard, P., Dalstein-Richier, L., 2015. Health and vitality assessment of two common pine species in the context of climate change in southern Europe. Environmental Research 137, 235e245. Silva, L.C., Go´mez-Guerrero, A., Doane, T.A., Horwath, W.R., 2015. Isotopic and nutritional evidence for species-and site-specific responses to N deposition and elevated CO2 in temperate forests. Journal of Geophysical Research: Biogeosciences 120 (6), 1110e1123. Singh, H., Pathak, P., Kumar, M., Raghubanshi, A.S., 2011. Carbon sequestration potential of Indo-Gangetic agroecosystem soils. Tropical Ecology 52 (2), 223e228. Stinziano, J.R., Way, D.A., 2014. Combined effects of rising [CO2] and temperature on boreal forests: growth, physiology and limitations 1. Botany 92 (6), 425e436. Sulman, B.N., Phillips, R.P., Oishi, A.C., Shevliakova, E., Pacala, S.W., 2014. Microbe-driven turnover offsets mineral-mediated storage of soil carbon under elevated CO2. Nature Climate Change 4 (12), 1099e1102. Thomas, D., Johannes, K., David, K., Ruediger, G., Ralf, K., 2016. Impacts of management and climate change on nitrate leaching in a forested karst area. Journal of Environmental Management 165, 243e252. Tian, Q., He, H., Cheng, W., Bai, Z., Wang, Y., Zhang, X., 2016. Factors controlling soil organic carbon stability along a temperate forest altitudinal gradient. Scientific Reports 6. Torn, M.S., Lapenis, A.G., Timofeev, A., Fischer, M.L., Babikov, B.V., Harden, J.W., 2002. Organic carbon and carbon isotopes in modern and 100-year-old-soil archives of the Russian steppe. Global Change Biology 8 (10), 941e953. Torn, M.S., Trumbore, S.E., Chadwick, O.A., Vitousek, P.M., Hendricks, D.M., 1997. Mineral control of soil organic carbon storage and turnover. Nature 389 (6647), 170e173. Treseder, K.K., Allen, M.F., 2000. Mycorrhizal fungi have a potential role in soil carbon storage under elevated CO2 and nitrogen deposition. New Phytologist 147 (1), 189e200. Urakawa, R., Shibata, H., Kuroiwa, M., Inagaki, Y., Tateno, R., Hishi, T., Fukuzawa, K., Hirai, K., Toda, H., Oyanagi, N., Nakata, M., Nakanishi, A., Fukushima, K., Enoki, T., Suwa, Y., 2014. Effects of freeze-thaw cycles resulting from winter climate change on soil nitrogen cycling in ten temperate forest ecosystems throughout the Japanese archipelago. Soil Biology & Biochemistry 74, 82e94. Vos, V.C., van Ruijven, J., Berg, M.P., Peeters, E.T., Berendse, F., 2013. Leaf litter quality drives litter mixing effects through complementary resource use among detritivores. Oecologia 173, 269e280. Von Arx, G., Pannatier, E.G., Thimonier, A., Rebetez, M., 2013. Microclimate in forests with varying leaf area index and soil moisture: potential implications for seedling establishment in a changing climate. Journal of Ecology 101 (5), 1201e1213. Wang, W., Peng, C., Kneeshaw, D.D., Larocque, G.R., Song, X., Zhou, X., 2012. Quantifying the effects of climate change and harvesting on carbon dynamics of boreal aspen and jack pine forests using the TRIPLEX-Management model. Forest Ecology and Management 281, 152e162. Wang, Y.P., Houlton, B.Z., 2009. Nitrogen constraints on terrestrial carbon uptake: implications for the global carbon-climate feedback. Geophysical Research Letters 36 (24), 2e5. Wieder, W.R., Cleveland, C.C., Smith, W.K., Todd-Brown, K., 2015. Future productivity and carbon storage limited by terrestrial nutrient availability. Nature Geoscience 8 (6), 441e444. Winsome, T., Silva, L.C., Scow, K.M., Doane, T.A., Powers, R.F., Horwath, W.R., 2017. Plant-microbe interactions regulate carbon and nitrogen accumulation in forest soils. Forest Ecology and Management 384, 415e423. Wurzburger, N., Hendrick, R.L., 2009. Plant litter chemistry and mycorrhizal roots promote a nitrogen feedback in a temperate forest. Journal of Ecology 97, 528e536. Xu, H.-J., Wang, X.-P., 2016. Effects of altered precipitation regimes on plant productivity in the arid region of northern China. Ecological Informatics 31, 137e146.

Chapter | Eight

Effects of Elevated CO2 in the Atmosphere on Soil C and N Turnover

Yakov Kuzyakov*, William R. Horwathx, Maxim Dorodnikov*, Evgenia Blagodatskaya*

*Department of Soil Science of Temperate Ecosystems, Department of Agricultural Soil Science, University of Goëttingen, Goëttingen, Germany; xDepartment Land, Air and Water Resources, University of California, Davis, CA, United States

INTRODUCTION Global change includes various components: (1) change of climate (temperature, precipitation) as well as frequency of extreme events (droughts, floods, heat and freeze waves, hurricanes and tornado); (2) modification of atmosphere (UV radiation; concentrations of CO2, CH4, N2O, O3, water vapor, VOC, etc.); (3) land use changes (forests clear-cut, wetlands drying, plowing); (4) soil degradation and decrease of soil fertility (erosion, salinization, nutrient losses); (5) acceleration of element’s cycling, both of nutrients (N, P, K, S, Ca, Mg) and ballast elements (Na, Si, Cl, Al, Se), but also toxic elements (mainly heavy metals); (6) human population growth with acceleration of urbanization and migration; (7) depletion of nonrenewable resources such as agricultural land, fresh and groundwater; fossil fuels, minerals); (8) chemical and organic pollution (eutrophication, pesticides, heavy metals, radionuclides, acid deposition, polyaromatic hydrocarbons); (9) decrease of biodiversity and genetic heterogeneity. From these global change components, the increase of the atmospheric CO2 level started to be worried firstly by Arrhenius (1896) who recognized a CO2 as the greenhouse gas. The main reason for our concern about atmospheric CO2 concentration is its direct relationship to the global temperature (Fig. 8.1). This very close connection between atmospheric CO2 and global temperature is proven at least for the last 500 k years (Barnola et al., 1991; Petit et al., 1999), but there are still intensive debates on what is the driver and what is Climate Change Impacts on Soil Processes and Ecosystem Properties. https://doi.org/10.1016/B978-0-444-63865-6.00008-9 Copyright © 2018 Elsevier B.V. All rights reserved.

207

208

Effects of Elevated CO2 in the Atmosphere CO2 concentration in the atmosphere and temperature Antartica Ice-core (Vostok)

CO2

Temperature

FIGURE 8.1 CO2 concentration in the atmosphere and temperature (Barnola et al., 1991; Petit et al., 1999).

the consequence. Based on the CO2 concentration changes slightly preceding the global temperature (Fig. 8.1), we can conclude that CO2 is the driver. Over the millennia, the atmospheric CO2 concentration remained below 300 ppm, and varied in the last 500 k years between w180 ppm during the ice ages and w280 ppm during the warm periods (Fig. 8.1). This was also the case during the last glaciation and the whole Holocene, but not during the Anthropocene. In the last 150e200 years the CO2 concentration in the atmosphere started to increase steadily. Year 2013 was the first year in the mankind history, when the CO2 concentration reached 400 ppm, and it is continuously increasing now. Therefore, it is crucial to understand how the elevated CO2 concentration in the atmosphere will affect processes in soildas a crucial part of each ecosystem. This is the main aim of this chapter, with more focus on soileplantemicrobial interactions and consequences for the C cycling. Despite continuous increase in atmospheric CO2 concentration, the rate of such an increase remarkably stabilized during the last 3e4 years indicating biosphere adaptation. Possible microbial mechanisms responsible for such an adaptation are discussed in this chapter.

APPROACHES TO INVESTIGATE INDIRECT EFFECTS OF ELEVATED CO2 CONCENTRATION ON SOIL PROCESSES Generally two groups of methods are used to investigate the effects of elevated CO2 concentration on ecosystems. First groupdmethods with controlled conditions, e.g., the leaf chambers and climate chambers, where the CO2 concentration is tested from ambient w400 ppm up to few percentage. These approaches are mainly focused to investigate metabolic and physiological mechanisms of changes at the level of few leafs or individual plants. They

Approaches to Investigate Indirect Effects

209

are not really suitable to investigate the response of plant communities and soil processes. In most cases these approaches do not consider the processes in roots and soil per se, and so, are not suitable to make any conclusions at the ecosystem level. The second group of methods considers multiple plants grown in soil and on the (small) ecosystem level under field conditions. This group includes (Table 8.1) natural CO2 springs, solar domes, open top chambers, and free air carbon dioxide enrichment (FACE) experiments. It is not our aim here to describe in details these approaches, and therefore, we just shortly mention their advantages and shortcomings (Table 8.1). For more details, see the extended reviews (McLeod and Long, 1999; Amthor, 2001). The main approach to assess the effects of elevated CO2 concentration on processes in soils and at ecosystem scale are FACE experiments because they have the lowest disturbance and the soil remains as a part of the ecosystem.

Table 8.1 Four Main Approaches to Investigate Effects of Elevated CO2 Concentration on Processes in PlanteSoil system Approaches

Advantages

Natural CO2 springs

l l

Natural CO2 source No costs

Shortcomingsa l l l l l

Solar domes

l

l

High CO2 enrichment is possible CO2 with shifted d13C

l l l l l

Open top chambers

l

l l

High CO2 enrichment is possible “Low” costs CO2 with shifted d13C

l l l l l

Free air carbon dioxide enrichment experiments

l l l l l

l

a

Ecosystem processes Low disturbance Long-term studies Medium scale Energy balance and gas exchange CO2 with shifted d13C

l

l

Few places around the world Azonal vegetation Very fast CO2 dilution Toxic gases Extreme soil conditions: l High soil temperature l High salinization Small scale (no upscaling) No ecosystem processes Mainly short term studies Higher temperature Higher humidity Small scale (no upscaling) No ecosystem processes Mainly short-term studies Higher temperature Higher humidity High costs of long-term CO2 maintenance High CO2 enrichment is impossible

Shortcomings are presented compared with natural conditions: ambient CO2 concentration (350e400 ppm) without any equipment.

210

Effects of Elevated CO2 in the Atmosphere

Hypotheses Considering the whole ecosystem (plantesoilemicroorganisms) and based on the classical Le-Chatelier’s principle, we should actually hypothesize that the increase of CO2 concentration in the atmosphere will lead to the maximal changes aboveground and the effects will subsequently decrease from leafs to shoots, roots, mycorrhiza, rhizosphere microorganisms, nonrhizosphere microorganisms and to soil organic matter (SOM) (Fig. 8.2). Please note, however, that this approach does not consider the interactions between the ecosystem components and the limitations and that the increase of CO2 concentration in the atmosphere will lead to stronger limitations of deliveries (supplies) from soildhere mainly the nutrients. Therefore, considering the shift of the limitations from the aboveground (atmospheric CO2 concentration) to the belowground (nutrients), we hypothesize that the increase of CO2 concentration in the atmosphere will lead to intensification of processes in soil with the main aim to overcome the nutrient limitation. This general hypothesis can be specified that nutrient mobilization in soil will be intensified (1) by abiotic processesdmainly chemical weathering, e.g., by release of more acidic agents by (1a) roots and (1b) microorganisms and (2) by biotic processes: (2a) microbial turnover and (2b) extracellular enzyme activities catalyzing decomposition of organic matter in soil and litter, thereby releasing nutrients.

--gaseous fluxes P potential priming effects R transport of refixed root-CO2 SOM shades: gradient of complexity

CO2

Interactions

Expected increase

Hypothesis

Pools and fluxes in the plant-soil system

Bahn et al. 2010 New Phytologist

FIGURE 8.2 Pools and fluxes in the pantesoil system. (Bahn et al., 2010, changed)

Results and Discussion

211

RESULTS AND DISCUSSION Direct and Indirect Effects of Elevated Atmospheric CO2 All effects (including effects of elevated CO2 concentration on soileplante microbial interactions) can be direct or indirect. Generally, it is tacitly assumed that direct effects are much stronger compared with indirect. However, this is not always the case for soils. To evaluate the direct effects of elevated CO2 concentration on soil processes, the CO2 concentrations in atmosphere and in soil should be compared. The CO2 concentration in soil is usually one or two orders of magnitude higher (5000 to 30,000 ppm, e.g., Pausch and Kuzyakov, 2012) than that in the atmosphere and strongly increases with depth. Also the variation of the CO2 concentration in soil (up to 10,000 ppm over short periods) is by far higher than the total CO2 concentration in the atmosphere. Consequently, the “small” increase of CO2 concentration in the atmosphere of “just” 100e200 ppm does not directly affect the CO2 concentration in soil at all. Even the CO2 gradient between soil and the atmosphere will not be affected. Therefore, elevated CO2 concentration in the atmosphere and even the expected increase of additional 50e100 ppm will not affect any soil process directly. Consequently, indirect effects are of much higher importance for soils. However, the investigation of indirect effects is much more difficult because they are based on various multilevel interactions: between the pools of elements, limiting factors, organisms, fluxes, etc.

Examples of Surprising and Unexpected Results Based on Interactions A good example of such interspecies interactions and partly unexpected consequences for plant growth was suggested for CeN rootemicrobial interactions (Fig. 8.3; Zak et al., 2000): (1) Plant production and root C input increase under elevated CO2 and so, (2) more C will be added to soil. This leads to (3) higher Interactions between plant and microbial activity,

CO2 and N availability by elevated CO2

Zak et al. 2000 New Phytologist

Tight interactions: roots microorganisms Competition for N Kuzyakov & Xu 2013 New Phytologist

FIGURE 8.3 Interactions between plant and microbial activity, and N availability by elevated CO2. (Zak et al. 2000, modified).

212


availability of organic substances for microorganisms, and so, to (4) microbial biomass growth with (5) higher N demand for microorganismsdincreased N immobilization. In turn, this leads to (6) less plant available N in soil, and so, to stronger competition between microorganisms and plants for N (Kuzyakov and Xu, 2013), resulting in N limitation for plants. Therefore, after the chain of these processes, plant growth and consequently the C release into the soil will be limited not by the CO2 concentration in the atmosphere but by mineral N in soil. Thus, despite removal of CO2 limitation under elevated CO2, plant productivity may not increase because of decreasing plant available N due to microbial N immobilization (Fig. 8.3). Another interesting and actually surprising result of elevated CO2 is the formation of soil aggregates (Rillig et al., 2001). At the first glance, there is no connection between the CO2 concentration in soil and atmosphere with the formation and stability of soil aggregates. However, because of the higher C input into the soil and increased N and P limitations, the mycorrhization of roots is stimulated (Godbold and Berntson, 1997). All mycorrhizal fungi release carbohydrates and proteins in soil, and some of arbuscular mycorrhiza, e.g., Glomus intraradices release glucoproteins such as glomalin having sticky properties, and so, lead to formation of larger and more stable aggregates. So, indirect effects of elevated CO2 on mycorrhizal production of gluing glucoproteins form and stabilize aggregates (Rillig and Mummey, 2006) and consequently improve water permeability and decrease erosion.

The Well-Known and Most Unknown Effects of Elevated CO2 Concentration on Belowground Processes Generally, it is well known that the plant biomass production increases under elevated CO2 for about 13%e20% (De Graaff et al., 2006) (but up to 200% for some crops, Rogers et al., 1994) compared with ambient. For the processes in soil, it is important that root growth increases stronger than that of aboveground biomass (Rogers et al., 1994). This includes the root growth rates and the amount of roots remaining in soil for longer periods (Ref. . ). However, not only the productivity butalso root morphology changes under elevated CO2: more fine and secondary roots will be produced (Pregitzer et al., 1995, 2000) because this is necessary for efficient acquisition of nutrients getting limiting under elevated CO2 (Luo et al., 2004). This also leads to increase of rhizodeposition (including exudation) as one of the mechanisms accelerating microbial mobilization of nutrients in the rhizosphere (Kuzyakov, 2002; Cheng ; Phillips . ; .). These fine roots as well as increased rhizodeposition stimulate CO2 production; but this CO2 is plant derived or originate from rhizomicrobial respiration and does not contribute to the increase of CO2 in the atmosphere (Kuzyakov, 2006). Besides the morphology changes, the quality of litter and root residues will be shifted toward higher C/N ratio and increased cellulose and lignin content


213

Effect of elevated CO2 on leaf litter quality

Norby et al. 2001 Oecologia

N content

lignin content

Weighted mean response ratio and 95% confidence interval The number to the right is the number of observations. If the confidence interval includes a response ratio of 1, the effect is not significant

Decrease of litter quality decrease of decomposition rates in soil

FIGURE 8.4 Effect of elevated CO2 on leaf litter quality. (Norby et al., 2001, modified).

(Fig. 8.4) (Norby et al., 2001; Norby and Jackson, 2000; Cotrufo et al., 2005). These shifts in the plant tissue quality decrease the decomposition rates and consequently prolong the residence time of litter in soil (Cotrufo et al., 2005). The most unknown effects of elevated CO2 are connected with processes in soil indirectly driven by plants. This includes effects on (1) Microorganisms (including mycorrhiza): composition, activity, growth, functions; (2) Extracellular enzymes: composition and activities; (3) SOM: content, composition of pools, turnover, allocation within soil matrix.

Effects of Elevated CO2 Concentration on Microbial Biomass and Functions Higher input of available C by roots into the soil stimulates microbial growth and biomass (Table 8.2, Zak et al., 2000). Despite average increase of microbial biomass for w16% compared with the soil under ambient CO2, the variation (SD) of the results is very high: up to 86%. This clearly shows that the final

Table 8.2 Change of Microbial Biomass and Microbial Respiration Under Elevated CO2 (Zak et al., 2000, Changed) Plants

Microbial Biomass

Microbial CO2

Graminoid

þ17 86

þ34 35

Herbaceous

þ29 29

þ34 19

Woody

þ19 46

þ20 23

Average

þ22

þ29

All data are presented as % changes under elevated CO2 compared with ambient CO2 (SD).

214


effects depend on the specific soil conditions, as partly mentioned aboved effects of N limitations. The increase of CO2 released by microbial respiration is higher than the microbial biomass and the variation is much less (Table 8.2). Consequently, we can conclude that elevated CO2 increases the fluxdmicrobial respirationdmore strongly than the pooldmicrobial biomass. Other important fluxes, reflecting microbial functions, also increased very strongly: soil respiration increased by 45%, microbial N immobilization by 93%, and net N mineralization by 44% compared with ambient CO2 (Zak et al., 2000). So, despite very high variation between the studies, all of them show that the most microbial processes will be accelerated under elevated CO2. Easily available C released by roots, stimulates microbial growth. Using the approach of substrate-induced growth respiration, Dorodnikov et al. (2009a,b) showed increase of specific microbial growth rates in bulk soil and in three aggregate size fractions of summer wheat ecosystem exposed to FACE. This increase for about 10% per hour, means that after about 7 h, the microbial biomass under elevated CO2 can be doubled compared with the soil under ambient CO2. Consequently, the increase of the pools (i.e., microbial biomass) is much less than that of the fluxes (i.e., microbial growth rates). The accelerated specific microbial growth rates were confirmed for three FACE experiments on different plant species with a clear linear relationship between growth rates and the CO2 increase (Blagodatskaya et al., 2010). An evidence that elevated CO2 effect was caused by altered quality and quantity of rhizodeposits is supported by the two times higher increase in specific microbial growth rates in the rhizosphere compared with the soil distant from the roots (Blagodatskaya et al., 2010). Remarkably, microbial response to elevated CO2 was much stronger when growth was induced by simple (glucose) versus complex (yeast extract) substrates. An increased C input in the form of rhizodeposits with a higher C/N ratio under elevated versus ambient CO2 (Cotrufo et al., 1994, 2005) decreases species diversity and selects microorganisms with lower auxotrophic requirements, i.e., those capable of rapid growth on energy-rich, simple substrates (Freeman et al., 2004). Thus, higher C input from plants is compensated by its faster turnover due to accelerated microbial growth. Such a faster microbial growth and turnover under elevated CO2 can partly explain the stabilization of CO2 increase rates in the atmosphere. An increase in microbial specific growth rates due to elevated CO2 was much stronger under N limitation, while N enrichment smoothed the effect of elevated CO2 on microbial growth (Blagodatskaya et al., 2010). Thus, the effect of elevated CO2 (faster growth) can be counterbalanced by N fertilization. This calls for the studies on dual (direct and indirect) N effect on plante microbial interactions in the rhizosphere under elevated CO2. N addition directly increases microbial C use efficiency and indirectly increases fine roots production thus, it boosts plantemicrobial competition for N in long term. A consequence of both direct and indirect effects of N is domination of more efficient, but slow-growing microorganisms are able to mine for N from SOM by production of extracellular enzymes.


215

Effects of Elevated CO2 on Soil Extracellular Enzymes Enzymes reflect activity of microorganisms for specific functions, nutrients, and element cycles. Again, strong evidence that elevated CO2 effects on microorganisms is mainly mediated by plants was obtained by simulation of root exudation with glucose input. Microorganisms activated by glucose doubled activity of hydrolytic enzymes under elevated versus ambient CO2 (Dorodnikov et al., 2009a,b). This confirmed accelerated C turnover under elevated CO2 in the rhizosphere. Remarkably, no CO2 effect on enzyme activity was detected in nonactivated soil. Elevated CO2 mainly stimulated enzymes related to N cycle and fungal activity (i.e., chitinase) and this effect was most pronounced in soil macroaggregates. This indicates a larger allocation of C from increased lignin and cellulose content (Cotrufo et al., 1994) to macroaggregates and increasing role of fungi (main decomposers of lignin and cellulose, Kirk and Farrel, 1987) to litter decomposition under elevated CO2. This is a demonstration of a possible mechanism: elevated CO2dincreased content of lignin in plantsdfungi produce more extracellular enzymes for N acquisition from plant and form microbial residuesdbut recalcitrant SOM is unaffected.

Effects of Elevated CO2 on Soil Organic Matter SOM is one of the most stable C pools in ecosystems, and consequently, its changes are usually small compared with the most other pools. So, the most FACE experiments (ongoing maximal w two decades) have only marginal SOM changes after years or even decades of elevated CO2 fertilization, especially when no additional nutrients input has been done (De Graaf et al., 2006). Even the contents of the SOM pools allocated between or inside differently sized aggregates as well as associated with soil minerals remain nearly constant (Dorodnikov et al., 2011). Based on the review of 59 studies, De Graaff et al. (2006) concluded that SOM increases only for 1.2%e2.2% of the stock year1 compared with that under ambient CO2. Therefore, we can conclude again that the C pools remain comparatively constant under elevated CO2 and its effect on the stable pools, like SOM, is hardly visible. Despite absence of the changes of SOM stocks, the SOM turnover in soils is sustained and could be altered under elevated CO2. To make the changes in SOM visible, the principal of tracing with the use of CO2 with distinct isotope characteristics, i.e., d13C, has been widely applied. Thus, the fumigated CO2 (e.g., in FACE studies) is usually derived from fossil fuel, which is depleted in 13C (d13C varies from 35& to 50&) compared with atmospheric CO2 (d13C 8&) (Hungate et al., 1996; Van Kessel et al., 2000). Plants grown in the elevated CO2 atmosphere are depleted in d13C and, as the litter and rhizodeposits from these plants decompose and become incorporated into the SOM, the soil d13C will decrease. The contribution of the new FACE-derived C to total SOM can be calculated based on the d13C of the SOM under ambient conditions and the period after initiating CO2 enrichment (Dorodnikov et al., 2008). Across many FACE studies the period of C occurrence in SOM (the so-called

216


mean residence time) varies from 3 to 4 years to 40e50 (Dorodnikov et al., 2008, 2011) and more than 100 years (Crow et al., 2009), and the increase of the SOM turnover rates under elevated CO2 is seemingly an established phenomenon (Thaysen et al., 2017). According to this, we may expect the acceleration of SOM turnover and even decomposition of relatively inert SOM pools in the future. Dissolved organic matter (DOM) represents an important constituent pool of the total SOM. Changes in DOM could help in the prediction of SOM transformations based on high mobility and availability of dissolved C fractions for decomposition. The observed intensification of SOM decomposition under elevated CO2 as mentioned above gets further confirmation in the increasing DOM pools in different ecosystems. For the forests (Phillips et al., 2011), the annual increase of dissolved organic inputs into soil comprised 50%, and for the modeled grassland communities DOC pool increased from 70% to 110% from the ambient level (Jones et al., 1998). These results imply that elevated atmospheric CO2 concentrations will have major impacts on soil food chains.

Priming Effects Under Elevated CO2 Increased rhizodeposition under elevated CO2 requires attention due to possible priming effect, i.e., accelerated decomposition of SOM caused by the input of labile C. This is especially important considering interactive effect of CO2 and N (see above). N input decreases microbial growth rates but increases growth efficiency indicating domination of K strategists able to produce enzymes decomposing recalcitrant SOM (Schimel and Schaffer, 2012). Stronger priming is expected, therefore in the absence of N limitation. In natural ecosystems, possible consequence of elevated CO2 is a shift in plant community due to invasion of plants with strong ability for N acquisition. Fast N uptake by strong competitors forces microorganisms to mineralize N from SOM, i.e., causes soil priming effect, which will be more pronounced under N limitation. In contrast, ambient plant community mitigate N limitation in the rhizosphere by release of more available C by roots to stimulate microorganisms for fast turnover and release of N from microbial biomass (Blagodatskaya et al., 2014).

CONCLUSIONS Considering the much higher CO2 concentration in soil compared with that in the atmosphere, we can clearly conclude that all the effects of elevated CO2 on soil processes are indirect effects: they are based on interactions between C input by plants into the soil, microbial biomass and activities, limitations by N (and probably other nutrients), and changes in water regimes. These indirect effects may be even larger than the direct effects of elevated CO2 expected for the vegetation. Most of these indirect effects are connected with the changes of the factor limiting plant and microbial growth: on the interactions between available C and N in soil.

References 217 Based on the broad range of processes, we can conclude that the effects of elevated CO2 on fluxes are much larger than on the pools. This was shown by comparison of microbial biomass and microbial functions, SOM and CO2 efflux from soil, DOM content and production, etc. In the most cases the pools remained the same under elevated CO2 compared with the ambient. It means that higher plant C input belowground under elevated CO2 was compensated by faster decomposition. Therefore, we can generally state that elevated CO2 concentration in the atmosphere will have no (or very small) effects on the pools but will increase of fluxes: both of them leads to acceleration of biogeochemical cycles.

REFERENCES Amthor, J.S., 2001. Effects of atmospheric CO2 concentration on wheat yield: review of results from experiments using various approaches to control CO2 concentration. Field Crops Research 73, 1e34. Arrhenius, S., 1896. On the influence of carbonic acid in the air upon the temperature of the ground. Philosophical Magazine and Journal of Science Series 5 (41), 237e276. Bahn, M., Janssens, I.A., Reichstein, M., Smith, P., Trumbore, S.E., 2010. Soil respiration across scales: towards an integration of patterns and processes. New Phytologist 186, 292e296. Barnola, J.M., Pimienta, P., Raynaud, D., Korotkevich, Y.S., 1991. CO2-climate relationship as deducted from the Vostok ice core - a reexamination based on new measurements and on a reevaluation of the air dating. TELLUS Series B e Chemical and Physical Meteorology 43, 83e90. Blagodatskaya, E., Blagodatsky, S., Dorodnikov, M., Kuzyakov, Y., 2010. Elevated atmospheric CO2 increases microbial growth rates in soil: results of three CO2 enrichment experiments. Global Change Biology 16, 836e848. Blagodatskaya, E., Littschwager, J., Lauerer, M., Kuzyakov, Y., 2014. Plant traits regulating N capture define microbial competition in the rhizosphere. European Journal of Soil Biology 61, 41e48. Cotrufo, M.F., Ineson, P., Rowland, A.P., 1994. Decomposition of tree leaf litters grown unders elevated CO2 e effect of litter quality. Plant and Soil 163, 121e130. Cotrufo, M.F., De Angelis, P., Polle, A., 2005. Leaf litter production and decomposition in a poplar short-rotation coppice exposed to free air CO2 enrichment (POPFACE). Global Change Biology 11, 971e982. Crow, S.E., Lajtha, K., Filley, T.R., Swanston, C.W., Bowden, R.D., Caldwell, B.A., 2009. Sources of plant-derived carbon and stability of organic matter in soil: implications for global change. Global Change Biology 15, 2003e2019. De Graaff, M.A., van Groenigen, K.J., Six, J., Hungate, B., van Kessel, C., 2006. Interactions between plant growth and soil nutrient cycling under elevated CO2: a meta-analysis. Global Change Biology 12, 2077e2091. Dorodnikov, M., Fangmeier, A., Giesemann, A., Weigel, H.J., Kuzyakov, Y., 2008. Thermal stability of soil organic matter pools and their turnover times at two levels of N fertilization calculated by d13C under elevated CO2. Isotopes in Health and Environmental Studies 44 (4), 365e376. Dorodnikov, M., Blagodatskaya, E., Blagodatsky, S., Fangmeier, A., Kuzyakov, Y., 2009a. Stimulation of r- vs. K- selected microorganisms by elevated atmospheric CO2 depends on soil aggregate size. FEMS Microbial Ecology 69, 43e52. Dorodnikov, M., Blagodatskaya, E., Blagodatsky, S., Marhan, S., Fangmeier, A., Kuzyakov, Y., 2009b. Stimulation of microbial extracellular enzyme activities by elevated CO2 depends on aggregate size. Global Change Biology 15 (6), 1603e1614.

218


Dorodnikov, M., Kuzyakov, Y., Fangmeier, A., Wiesenberg, G.L.B., 2011. C and N in soil organic matter density fractions under elevated CO2: turnover vs. stabilization. Soil Biology & Biochemistry 43 (3), 579e589. Freeman, et al., 2004. Godbold, D.L., Berntson, G.M., 1997. Elevated atmospheric CO2 concentration changes ectomycorrhizal morphotype assemblages in Betula papyrifera. Tree Physiology 17, 347e350. Hungate, B.A., Jackson, R.B., Field, C.B., Chapin III, F.C., 1996. Detecting changes in soil carbon in CO2 enrichment experiments. Plant Soil 187, 135e145. Jones, T.H., Thompson, L.J., Lawton, J.H., Bezemer, T.M., Bardgett, R.D., Blackburn, T.M., Bruce, K.D., Cannon, P.F., Hall, G.S., Hartley, S.E., et al., 1998. Impacts of rising atmospheric carbon dioxide on model terrestrial ecosystems. Science 280 (5362), 441e443. Kirk, Farrel, 1987. Kuzyakov, Y., 2002. Review: factors affecting rhizosphere priming effects. Journal of Plant Nutrition and Soil Science 165 (4), 382e396. Kuzyakov, Y., 2006. Sources of CO2 efflux from soil and review of partitioning methods. Soil Biology & Biochemistry 38, 425e448. Kuzyakov, Y., Xu, X., 2013. Competition between roots and microorganisms for N: mechanisms and ecological relevance. New Phytologist 198, 656e669. Luo, Y., Su, B., Currie, W.S., Dukes, J.S., Finzi, A.C., Hartwig, U., Hungate, B., McMurtrie, R.E., Oren, R., Parton, W.J., et al., 2004. Progressive nitrogen limitation of ecosystem responses to rising atmospheric carbon dioxide. Bioscience 54, 731e739. McLeod, A.R., Long, S.P., 1999. Free-air carbon dioxide enrichment (FACE) in global change research: a review. Advances in Ecological Research 28, 1e56. Norby, R.J., Jackson, R.B., 2000. Root dynamics and global change: seeking an ecosystem perspective. New Phytologist 147, 3e12. Norby, R.J., Cotrufo, M.F., Ineson, P., O’Neill, E.G., Canadell, J.G., 2001. Elevated CO2, litter chemistry, and decomposition: a synthesis. Oecologia 127, 153e165. Pausch, J., Kuzyakov, Y., 2012. Soil organic carbon decomposition from recently added and older sources estimated by d13C values of CO2 and organic matter. Soil Biology & Biochemistry 55, 40e47. Petit, J.R., Jouzel, J., Raynaud, D., Barkov, N.I., Barnola, J.M., Basile, I., Bender, M., Chappellaz, J., Davis, M., Delaygue, G., et al., 1999. Climate and atmospheric history of the past 420,000 years from the Vostok ice core, Antarctica. Nature 399 (6735), 429e436. Phillips, R.P., Finzi, A.C., Bernhardt, E.S., 2011. Enhanced root exudation induces microbial feedbacks to N cycling in a pine forest under long-term CO2 fumigation. Ecology Letters 14, 187e194. Pregitzer, K.S., Zak, D.R., Curtis, P.S., Kubiske, M.E., Teeri, J.A., Vogel, C.S., 1995. Atmospheric CO2, soil-nitrogen and turnover of fine roots. New Phytologist 129, 579e585. Pregitzer, K.S., Zak, D.R., Maziasz, J., DeForest, J., Curtis, P.S., Lussenhop, J., 2000. Interactive effects of atmospheric CO2 and soil-N availability on fine roots of Populus tremuloides. Ecological Applications 10, 18e33. Rillig, M.C., Mummey, D.L., 2006. Mycorrhizas and soil structure. New Phytologist 171, 41e53. Rillig, M.C., Wright, S.F., Kimball, B.A., Pinter, P.J., Wall, G.W., Ottman, M.J., Leavitt, S.W., 2001. E-carbon dioxide and irrigation effects on water stable aggregates in a Sorghum field: A possible role for arbuscular mycorrhizal fungi. Global Change Biology 7, 333e337. Rogers, H.H., Runion, G.B., Krupa, S.V., 1994. Plant-responses to atmospheric CO2 enrichment with emphasis on roots and the rhizosphere. Environmental Pollution 83, 155e189. Schimel, J.P., Schaeffer, S.M., 2012. Microbial control over carbon cycling in soil. Frontiers in Microbiology 3, 348. https://doi.org/10.3389/fmicb.2012.00348. Thaysen, E.M., Reinsch, S., Larsen, K.S., Ambus, P., 2017. Decrease in heathland soil labile organic carbon under future atmospheric and climatic conditions. Biogeochemistry 133, 17e36.

References 219 Van Kessel, C., Nitschelm, J., Horwath, W.R., Harris, D., Walley, F., Luescher, A., Hartwig, U., 2000. Carbon-13 input and turnover in a pasture soil exposed to longterm elevated atmospheric CO2. Global Change Biology 6 (1), 123e135. Zak, et al., 2000. Elevated atmospheric CO2, fine roots and the response of soil microorganisms. New Phytologist 147, 201e222.


Index ‘Note: Page numbers followed by “f ” indicate figures, “t” indicate tables and “b” indicate boxes.’

A Acidobacteria, 114e115 Acrisols, 187e188 Actinobacteria, 113e114 Agricultural diversification, 19 Alphaproteobacteria, 114e115 Amazon, 10e12 American Southwest, 19 Ammonia monooxygenases, 121 Ammonia oxidation pathways acetylene, 135e136 Ancient agricultural systems, 3, 12 soil degradation, 14e16 sustainability, 13e14 Ancient agricultural terraces, Colca Valley, Peru, 10e12, 11f Ancient civilizations, 3, 8 Ancient cultures, 17e20 Ancient sites advantages, 5e7 agricultural degradation, 5, 6t benefits, 5e7 deep time, 4e5 limitations, 5 soil properties, 5, 6t Andes Mountains of South America, 14 Andisols, 98e100 Animal domestication, South America, 10e12 Animal manures sequester soil organic carbon composted poultry manure, 65 conservation tillage impact, 66e67 cover crops and crop rotation effects, 67e68 human waste, 65 Rothamsted Experiment, 65 soil amendment, 65 Anthropocene, 33e34 Anthropogenic interactions, 17 Apparent thermal acclimation, 114e115 Arbuscular mycorrhizal (AM) fungal abundance, 113

Archaeological record, 4e5 advantages, 5e7 interpretations, 5 soil productivity, 5 Archaeology, 3, 4f Assigning causality, 18

B Batch sorption, 104 Bicarbonate salts, 94 Biodiversity, 30 hotspots, 36e37 Biogeochemical mechanisms, 43be44b Biomass, 190 Black box, 112 Bolivian Amazon, 10e12

C Calcisols, 187e188 Carbon sequestration, 32e33 Carthaginians, 9e10 Chemodenitrification, 146e147 Chronosequence, 100e102, 101f Climate amelioration, 17e18 Climate friendly agricultural systems, 72e73 Climatic fluctuations, 1e2 Colca Valley, Peru, 10e12, 11f, 14 Comammox, 144e145 Conservation tillage, 66e67 Copper Age settlement, 3 Cover crops, 8e9, 9f, 67e68 Crop rotation effects, 67e68

D Dark Earth, 10e12, 12f Degradational effects, 20 Drought, 115e116

E East Asian techniques, 16e17 Easter Island, 15 EcoPlate, 121e122 Egyptian civilization, 9e10

221

222

Index

Elevated CO2 atmosphere modification, 207, 208f biodiversity and genetic heterogeneity, 207 chemical and organic pollution, 207 climate changes, 207 direct and indirect effects, 211 element’s cycling acceleration, 207 examples, 212 free air carbon dioxide enrichment (FACE), 209e210 human population growth, 207 hypotheses, 210e211, 210f land use changes, 207 microbial biomass, 213t, 214e215 nonrenewable resources depletion, 207 priming effects, 216e217 soil extracellular enzymes, 215 soil organic matter (SOM), 210, 215e216 soil processes indirect effects, 208e211, 209t unknown effects, 212e213 well-known effects, 212e213 Escherichia coli, 119

F Feammox, 144 Fertilization, 147e155 fertilizer application timing, 152e153 fertilizer efficiency enhancers, 153e155 fertilizer placement, 151e152 fertilizer rate, 148e149 fertilizer type, 149e151 mitigation multiple issues, 155 Flooding, 116e117 Forest ecosystems Acrisols, 187e188 biomass, 190 Calcisols, 187e188 capacity, 195e196 carbon sequestration, 193e195, 194f climate change, 187 CO2 mineralization, 197f dendrochronological series, 189 environmental services, 195e196 Ferralsols, 187e188 free-air CO2 enrichment (FACE) experiments, 189 global distribution of studies, 187e188, 187f interrelated factors networks, 189e190 Larix, 189e190 leaf area index (LAI), 189 Leptosols, 187e188 litter production, 191 Lixisols, 187e188

microbial processes, 198e201, 199f net primary productivity (NPP), 185, 188e193 Nitosols, 187e188 nutrient availability, 190e191 Picea abies (L.), 192 Picea marina, 198 pinion, 191e192 Pinus banksiana, 198 Pinus halepensis L., 186 priming effect, 188e189 Quercus suber, 197e198 rainfall, 192 soil texture, 196e198, 197f temperature and precipitation, 195 Free-air CO2 enrichment (FACE) experiments, 189 Fungal denitrification, 143

G Global reservoirs, 95t Green Revolution, 62e63

H HabereBosch process, 61e62 Harvest index (HI), 74e75, 74f Heterotrophic denitrification, 140e142 Heterotrophic nitrification, 133e135, 142e143 Hydrological shifts, 43be44b Hydroxylamine decomposition, 145e146

I Incubation experiments, 104 Induced magnetism, 3 Intercropping, 10e12 maize and squash, 10, 11f Irrigation, 156e158 flood irrigation, 156e157 furrow irrigations, 157 subirrigation, 157 subsurface drip irrigation, 157e158

K K-strategists/oligotrophic microbes, 117e118

L Larix, 189e190 Leaf area index (LAI), 189 Leptosols, 187e188 Lixisols, 187e188

M Mayan agriculture, 10 Microbial response traits, 119e121, 122t

Index Mineral assemblage, 96e102 Mineralogy, 93e95 Mineral stabilization, 102e103 Mormon Trail in Iowa, USA, 17, 17f Multimodel mean projections, 2, 2f

hydroxylamine decomposition, 132e133, 145e146 irrigation, 156e158 flood irrigation, 156e157 furrow irrigations, 157 subirrigation, 157 subsurface drip irrigation, 157e158 land management practices, 147 Nitrospira, 144e145 NO3- reduction to NH4+, 143e144 organic amendments, 160e162 Pseudomonas denitrificans, 142 soil production, 162e164 synthetic N fertilizers, 161 tillage, 158e160 conservation tillage, 159 N fertilizer management, 158e159

N Neolithic settlements, 1 Net primary productivity (NPP), 185 Nitosols, 187e188 Nitrification-coupled denitrification, 138 Nitrifier denitrification, 135e138 Nitrifier nitrification, 135 Nitrite oxidoreductase, 121 Nitrosomonas europaea, 135e136 Nitrospira, 144e145 Nitrous oxide (N2O) abiotic processes, 132e133 abiotic production, 145 ammonia oxidation pathways, 135e140 acetylene, 135e136 factors, 139e140, 140f land management practices, 137f metals, 139e140 nitrification-coupled denitrification, 138 nitrifier denitrification, 135e138 nitrifier nitrification, 135 Nitrosomonas europaea, 135e136 oxygen atoms (O), 136e137 oxygen availability, 139 biological processes, 133e135, 134f chemodenitrification, 132e133, 146e147 climate change, 162e164, 162f comammox, 144e145 cover crops, 160e162 emission sources, 131, 132f enzymatic processes, 132e133 feammox, 144 fertilization, 147e155 fertilizer application timing, 152e153 fertilizer efficiency enhancers, 153e155 fertilizer placement, 151e152 fertilizer rate, 148e149 fertilizer type, 149e151 mitigation multiple issues, 155 fungal denitrification, 143 greenhouse gas (GHG), 131, 133 heterotrophic denitrification, 140e142 factors, 141e142 multiple laboratory incubations, 141 heterotrophic nitrification, 133e135, 142e143

223

O Organic amendments, 160e162

P Paleoenvironmental reconstruction, 3 Paleozoic ice age, 32 Picea abies (L.), 192 Picea marina, 198 Pinus banksiana, 198 Pinus halepensis L., 186 Pinus uncinata, 189e190 Pseudomonas denitrificans, 142

Q Quercus suber, 197e198

R Resource-acquisition strategies, 30 Rice terraces, 8, 8f Rothamsted Experiment, 65

S Snowfall, 117 SOC. See Soil organic carbon (SOC) Soil carbon cycle andisols, 98e100 batch sorption, 104 bicarbonate salts, 94 bonding mechanisms, 102e103 carbonate parent materials, 94 chronosequence, 100e102, 101f climate, 96e102 CO2, 93e95 field and lab-based evidence, 104e107 fluxes, 95 global reservoirs, 95t incubation experiments, 104 inorganic C cycle, 93e95

224

Index

Soil carbon cycle (Continued) mineral assemblage, 96e102 mineralogy, 93e95 mineral stabilization, 102e103 organic C cycle, 96e102 organomineral bonding, 102e103 phyllosilicate phases, 102 physiochemical control, 96e100 reactive properties, 103t recent advances, 106e107 sizes, 95 SOC stocks and stability, 97, 99f soil-forming factors, 97, 98f soil mineralogical properties, 104e105 Strakhov diagram, 96e97, 96f weathering, 93e95 Soil chemistry, 3 Soil degradation Easter Island, 15 Mesopotamia, 14 Norse in Greenland, 16 “rat outbreak impact”, 15e16 Soil-forming factors, 97, 98f Soil knowledge/management, early civilizations, 7 Africa, 9e10 Americas ancient agricultural terraces, Colca Valley, Peru, 10e12, 11f intercropped maize and squash, 10, 11f Mayan agriculture, 10 terra preta, 10e12, 12f Asia, 7e8, 8f Europe, 8e9, 9f Soil microbial communities Acidobacteria, 114e115 Actinobacteria, 113e114 Alphaproteobacteria, 114e115 ammonia monooxygenases, 121 apparent thermal acclimation, 114e115 arbuscular mycorrhizal (AM) fungal abundance, 113 bacteroidetes, 113e114 black box, 112 C and N transformations, 121 climate change drivers, 116e117 climate change impacts research, 112e117 climate change prediction, 117e123, 120fe121f DGGE, 115e116 DNA-based approaches, 113 DNA-based sequencing methods, 114e115 drought, 115e116

EcoPlate, 121e122 ecosystem functioning, 111e112 elevated CO2, 112e113 Escherichia coli, 119 flooding, 116e117 fungal PLFA, 115e116 K-strategists/oligotrophic microbes, 117e118 microbial response traits, 119e121, 122t nitrite oxidoreductase, 121 oligotrophic microbial life-history strategies, 118 PCR-denaturing gradient gel electrophoresis (DGGE), 113 snowfall, 117 soil organisms, 112 warming, 114e115 Soil micromorphological analysis, 3 Soil mineralogical properties, 104e105 Soil organic carbon (SOC), 39e40 agricultural soil C sequestration rates, 68e70, 68te69t animal manures sequester soil organic carbon composted poultry manure, 65 conservation tillage impact, 66e67 cover crops and crop rotation effects, 67e68 human waste, 65 Rothamsted Experiment, 65 soil amendment, 65 atmospheric composition, 81e82 carbon balance and management factors, 63e64 management practices, effect of, 63e64, 64f soil productivity, 64e65 climate change, 78e79 cropland climate friendly agricultural systems, 72e73 cover crops, 73e74 crop yields over time, 75e76, 75f harvest index (HI), 74e75, 74f low-input management, 72e73 residue production over time, 75e76, 75f forest, 77 soil C sequestration rates, 68e70, 71t grasslands, 76e77 soil C sequestration rates, 68e70, 70t humanities reliance and impact fossil fuel, 62e63 Green Revolution, 62e63 HabereBosch process, 61e62

Index increasing human population, 61e62, 62f productive agriculture land, 61e62, 63f importance, 61 N requirement, 79e81, 79f research in, 82e83 stabilization, 61 wetlands, 77e78 soil C sequestration rates, 68e70, 72t Soil organisms, 112 Soil-plant-atmosphere (SPA) interactions, 31, 47e48 biogeochemical mechanisms, 43be44b biological associations, 43be44b bottom-up regulation of symbiotic associations interplant resource transfer, 39 plant diversity trajectory, 37e38, 37f predominant resource-acquisition strategies, 37e38, 37f root-microbe networks, 39 soil formation and carbon pools, 38 conservation and management opportunities, 45e46 boreal forest expansion, 46e47 global land area, 46e47, 46f hydrological shifts, 43be44b mathematical models, 42e44 simplifying complexity, 40f atmospheric CO2, 39e40 climate warming, 41 microorganisms, 39e40 soil carbon pool, 40e41 soil organic carbon, 39e40 soil and ecosystem development biodiversity hotspots, 36e37 edaphic and ecological trajectories, 36 global map of soil orders (USDANRCS), 34e36, 35f

225

human dependency, 36 vascular plant diversity and distribution, 34e36, 35f terrestrial life, 31 arbuscular mycorrhizal associations, 32e33 carbon sequestration, 32e33 glacial-interglacial climate oscillations, 33 Paleozoic ice age, 32 preindustrial fluctuations, 33e34 recent ecosystem, 33e34, 34b temperatures, CO2, and biodiversity curves, 31e32, 32f Soil texture, 196e198, 197f Strakhov diagram, 96e97, 96f Sustainability, 13e14

T Terra preta, 10e12, 12f Terrestrial ecosystems, 33e34 Terrestrial life, 31 arbuscular mycorrhizal associations, 32e33 carbon sequestration, 32e33 glacial-interglacial climate oscillations, 33 Paleozoic ice age, 32 preindustrial fluctuations, 33e34 recent ecosystem, 33e34, 34b temperatures, CO2, and biodiversity curves, 31e32, 32f Terrestrial organisms, 30 Tigris and Euphrates rivers, 13 Tillage, 158e160

V Vascular plant diversity/distribution, 34e36, 35f

W Warming, 114e115 Weathering, 93e95 Wetlands, 77e78
