Environmental Engineering and Management Journal

July 2013 Vol.12 No. 7

ISSN 1582 - 9596

Environmental Engineering and Management Journal An International Journal Editor-in-Chief:

Matei Macoveanu

Guest Editors:

Yaqian Zhao Ji-Dong Gu Xinmin Zhan

Managing Editor:

Maria Gavrilescu

RECENT ADVANCES IN WATER RESOURCE MANAGEMENT AND POLLUTION CONTROL: WITH SPECIAL FOCUS ON CHINA

“Gheorghe Asachi” Technical University of Iasi

Environmental Engineering and Management Journal

EcoAdvertising Offer

The Environmental Engineering and Management Journal encourages initiatives and actions concerning the improvement of education, research, marketing and management, in order to achieve sustainable development. This journal brings valuable opportunities for those offering products, technologies, services, educational programs or other related activities, creating thus a closer relation with the request of the market in the fields of environmental engineering, management and education. This journal address researchers, designers, academic staff, specialists with responsibilities in the field of environmental protection and management from government organizations (central and local administrations, environmental protection agencies) or from the private or public companies. Also, graduates of specialization courses or of the Environmental Engineering and Management profile, as well as other specialists may find in this journal a direct linkage between the offer and request of the market concerned with the protection of the environment and the administration of natural resources in the national and international context. The journal was conceived as a means to develop scientific and technical relationships between people who offer and request solutions for environmental protection and conservation of natural resources, creating thus the premises to enhance the transfer of technology and know-how, the confirmation and implementation of ecological products and services. Taking these aspects into consideration, we gladly welcome any persons or companies which correspond to the abovementioned purposes and objectives to use our journal to identify potential collaborators, thus contributing to the support of this selffinanced journal, which has a special section for eco-advertising.

Costs of advertising material publication I.

Color ad – whole page, inside 160 € Color ad – ½ page, inside 90 € Color ad – ¼ page, inside 50 € 2 Color ad – other format, inside 0.50 €/cm II. Black and white ad – whole page, inside 120 € Black and white ad – ½ page, inside 75 € Black and white ad – ¼ page, inside 45 € 2 Black and white ad – other format, inside 0.40 €/cm III. Covers III, IV 220 € IV. Advertising article 80 €/page At a constant publication of the advertising material, the following discounts will be offered:  3 publications 10%  6 publications 15%  12 publications 20% __________________________________________________________________________________ For subscriptions and orders please contact us at: O.A.I.M.D.D (Academic Organization for Environmental Engineering and Sustainable Development), 71 Mangeron Blvd., PO 10, Box 2111, 700050 Iasi, Romania, CIF: 10150285 Phone/Fax: 0040-232-271759 E-mail: [email protected] [email protected] In lei: BRD IASI, RO53BRDE240SV08296282400 In EURO: Romanian Bank for Development, Groupe Société Generale, Bucharest, Romania SWIFT Code: BRDEROBU Beneficiary: Iasi, Romania, RO44BRDE240SV09790262400 __________________________________________________________________________________ Department of Environmental Engineering and Management 73 Prof.Dr.docent Dimitrie Mangeron Street, 700050-Iasi, Romania Phone/Fax: 0040-232-271759 E-mail: [email protected] [email protected]

Environmental Engineering and Management Journal EDITORIAL BOARD Editor-in-Chief: Matei Macoveanu Gheorghe Asachi Technical University of Iasi, Romania

Managing Editor: Maria Gavrilescu Gheorghe Asachi Technical University of Iasi, Romania

SCIENTIFIC ADVISORY BOARD Ahmet Aktaş Akdeniz University, Antalya Turkey

Anca Duta Capra Transilvania University of Brasov Romania

Valentin Nedeff Vasile Alecsandri University of Bacau Romania

Maria Madalena dos Santos Alves University of Minho, Braga Portugal

Fabio Fava Alma Mater Studiorum University of Bologna Italy

Vadim I. Nedostup Academy of Science, Physico-Chemical Institute, Odessa, Ukraine

Abdeltif Amrane University of Rennes, ENSCR France

Eugenio Campos Ferreira University of Minho, Braga, Portugal

Alexandru Ozunu Babes-Bolyai University of Cluj-Napoca Romania

Ecaterina Andronescu University Polytechnica of Bucharest Romania

Cristian Fosalau Gheorghe Asachi Technical University of Iasi Romania

Yannis A. Phillis Technical University of Crete, Chania Greece

Robert Armon Technion-Israel Institute of Technology, Haifa Israel

Anton Friedl Vienna University of Technology Austria

Marcel Ionel Popa Gheorghe Asachi Technical University of Iasi Romania

Adisa Azapagic The University of Manchester United Kingdom

Anne Giroir Fendler University Claude Bernard Lyon 1 France

Marcel Popa Gheorghe Asachi Technical University of Iasi Romania

Hamidi Abdul Aziz Universiti Sains Malaysia, Penang Malaysia

Ion Giurma Gheorghe Asachi Technical University of Iasi Romania

Valentin I. Popa Gheorghe Asachi Technical University of Iasi Romania

Pranas Baltrenas Vilnius Gediminas Technical University Lithuania

Aurelian Gulea State University of Moldavia, Kishinew Republic of Moldavia

Tudor Prisecaru University Polytechnica of Bucharest Romania

Hans Bressers University of Twente, Enschede The Netherlands

Yuh-Shan Ho Peking University People's Republic of China

Gabriel-Lucian Radu Polytechnica University of Bucharest Romania

Han Brezet Delft University of Technology The Netherlands

Arjen Y. Hoekstra University of Twente, Enschede The Netherlands

Ákos Rédey Pannon University, Veszprém Hungary

Dan Cascaval Gheorghe Asachi Technical University of Iasi Romania

Nicolae Hurduc Gheorghe Asachi Technical University of Iasi Romania

Joop Schoonman Delft University of Technology The Netherlands

Alexandru Cecal Al.I. Cuza University of Iasi Romania

Ralf Isenmann Munich University of Applied Sciences Germany

Dan Scutaru Gheorghe Asachi Technical University of Iasi Romania

Aleg Cherp Central European University, Budapest Hungary

Marcel Istrate Gheorghe Asachi Technical University of Iasi Romania

Ilie Siminiceanu Gheorghe Asachi Technical University of Iasi Romania

Yusuf Chisti Massey University, Palmerston North New Zealand

Ravi Jain University of Pacific, Baun Hall Stockton United States of America

Bogdan C. Simionescu Gheorghe Asachi Technical University of Iasi Romania

Philippe Corvini University of Applied Sciences Northwestern Switzerland, Muttenz, Switzerland

Michael Søgaard Jørgensen Aalborg University Denmark

Florian Statescu Gheorghe Asachi Technical University of Iasi Romania

Andrew J. Daugulis Queen's University Kingston Canada

Nicolas Kalogerakis Technical University of Crete, Chania Greece

Vyacheslav N. Stepanov Academy of Sciences, Institute of Market and EconomicEcological Research, Odessa, Ukraine

Valeriu David Gheorghe Asachi Technical University of Iasi Romania

Gheorghe Lazaroiu University Polytechnica of Bucharest Romania

Carmen Teodosiu Gheorghe Asachi Technical University of Iasi Romania

Katerina Demnerova University of Prague Czech Republic

Thomas Lindhqvist International Institute for Industrial Environmental Economics, Lund University, Sweden

Saulius Vasarevicius Vilnius Gediminas Technical University Lithuania

Gheorghe Duca State University of Moldavia, Kishinew Republic of Moldavia

Andreas Paul Loibner University of Natural Resources and Life Sciences, Vienna, Austria

Angheluta Vadineanu The University of Bucharest Romania

Ioan Dediu Academy of Sciences, Institute of Ecology and Geography, Kishinev, Republic of Moldavia

Tudor Lupascu Academy of Sciences, Institute of Chemistry, Kishinev, Republic of Moldavia

Colin Webb The University of Manchester United Kingdom

Emil Dumitriu Gheorghe Asachi Technical University of Iasi Romania

Antonio Marzocchella University of Naples Federico II, Naples, Italy

Peter Wilderer Technical University Munich Germany

Jurek Duszczyk Delft University of Technology The Netherlands

José Mondéjar Jiménez University Castilla-La Mancha, Cuenca Spain

Petra Winzer Bergische University Wuppertal Germany

Shin' ichi Nakatsuji University of Hyogo Japan

Environmental Engineering and Management Journal Environmental Engineering and Management Journal is included and indexed in CABI Chemical Abstracts Service/SciFinder (ACS) (since 2002) EBSCO Database (since 2002) EVISA ICAAP (International Consortium for Advancement of Academic Publications) Index Copernicus Journal Master List (ICV/2010=16.20) Journal Citation Reports® (IF=1.117), (Environmental Sciences, Ranked 146 of 206), (5-Year Impact Factor: 0.970 Article Influence® Score: 0.085) MedSci ProQuest (since 2002) The National University Research Council (RO) Science Citation Index Expanded™ (Thomson ISI) SJR (SCImago Journal&Country Rank) (Environmental Sciences, Ranked 480 of 825, H=14, SJR index/2012 = 0.306, SNIP index/2012 = 0.76) SCOPUS (since 2008) Thomson ISI Master Journal List Web of Science® (Thomson ISI) (H=16)

Home page: http://omicron.ch.tuiasi.ro/EEMJ/ Full text: http://www.ecozone.ro

Editor-in-Chief: Matei Macoveanu, Iasi (RO) Managing Editor: Maria Gavrilescu, Iasi (RO) Gheorghe Asachi Technical University of Iasi Faculty of Chemical Engineering and Environmental Protection Department of Environmental Engineering and Management – Editorial and Production Office 73 Prof.Dr.docent Dimitrie Mangeron Street, 700050 Iasi, Romania Phone: +40-232-278680, ext. 2240 Fax: +40-232-271759 e-mail: eemjournal.at.yahoo.com, eem_journal.at.yahoo.com, eemjeditor.at.yahoo.com, eemj_editor.at.yahoo.com, eemjournal.at.gmail.com, eemj.editor.at.gmail.com, eemj.office.at.gmail.com Editorial production and secretariat: Camelia Smaranda Laura Carmen Apostol Raluca-Maria Hlihor Petronela Cozma Cristina Ghinea Isabela Simion Dana Luminiţa Sobariu . Published 12 issues per year, under the aegis of the “Gheorghe Asachi” Technical University of Iasi, Romania by EcoZone Publishing House of the Academic Organization for Environmental Engineering and Sustainable Development (OAIMDD), http://www.ecozone.ro Annual subscription rate 2012 (12 issues) Print: EUR EUR

350 per volume 40 per issue

Electronic: 300 per volume 35 per issue

Order directly to the Editorial Office 73 Prof.Dr.docent Dimitrie Mangeron Street, 700050 Iasi, Romania Phone/Fax: Fax: +40-232-271759 e-mail: mmac.at.ch.tuiasi.ro mgav_eemj.at.yahoo.com Electronic, full text: Order or purchase on-line at: www.ecozone.ro Bank account (EURO): Romanian Bank for Development, Groupe Societé Generale, Bucharest, Romania SWIFT Code: BRDEROBU Beneficiary: Iasi, Romania, RO44BRDE240SV09790262400 All rights reserved, including those of translation into foreign languages. No part of each issue may be reproduced in any form (photoprint, microfilm, or any other means) nor transmitted or translated without written permission from the publishers. Only single copies of contributions, or parts thereof, may be made for personal use. This journal was carefully produced in all its parts. Even so, authors, editors and publisher do not guarantee the information contained there to be free of errors. Registered names, trademarks etc. used in this journal, even when not marked as such, are not be considered unprotected by law.


July 2013, Vol.12, No. 7, 1309-1528

http://omicron.ch.tuiasi.ro/EEMJ/

“Gheorghe Asachi” Technical University of Iasi, Romania

CONTENTS ____________________________________________________________________________________________

Editorial A special issue on: RECENT ADVANCES IN WATER RESOURCE MANAGEMENT AND POLLUTION CONTROL: WITH SPECIAL FOCUS ON CHINA Yaqian Zhao, Jidong Gu, Xinmin Zhan …...………………………………………………………….

1309

Papers China’s water management – challenges and solutions Yanan Jiang, Faith Ka Shun Chan, Joseph Holden, Yaqian Zhao, Dabo Guan ………………..

1311

Radiation-induced decomposition and polymerization Weihua Sun, Lujun Chen, Jinping Tian, Jianlong Wang , Shijun He.........................................

1323

Effect of aeration rate on Domestic Wastewater treatment using an Intermittently Aerated Sequencing Batch Reactor (IASBR) technology Liam G. Henry, Liwen Xiao, Xinmin Zhan …………………………………………………………..

1329

Characterization of nitrification performance and microbial community in a MBBR and integrated GBBR-MBBR treating heavily polluted river water Xiangchun Quan, Linyun Gu, Yin Qian, Yuansheng Pei, Zhifeng Yang…………………………

1335

Effects of Fe(III) on dissimilatory ferric reduction, nitrogen and phosphorus removal in activated sludge process Ya’e Wang, Jie Li, Juanjuan Feng, Siyuan Zhai.........................................................................

1345

Whole cell bioreporter for the estimation of oil contamination Chuan Li, Dayi Zhang, Yizhi Song, Bo Jiang, Guanghe Li, Wei E. Huang ……………………...

1353

Control mode of rural nonpoint source pollution in Tai Lake Basin, China Yimin Zhang, Yuexiang Gao, Jingcheng Duan, Yang Yang, Chuang Zhou, Houhu Zhang……………………………………………………………………………………….........

1359

Enhanced nitrogen removal from reject water of municipal wastewater treatment plants using a novel excess activated sludge (EAS) based nitrification and denitrification process Yongzhe Yang, Xiaoxia Yang, Wei Li, Xinchao Guo...................................................................

1367

Aerobic sludge granulation for partial nitrification of ammonia-rich inorganic wastewater An-Jie Li, Xiao-Yan Li, Xiang-Chun Quan, Zhi-Feng Yang………………………………………..

1375

Adsorption of hexachlorocyclohexane by raw and surfactant modified meerschaum Shengke Yang, Wenke Wang, Yue Zhao, Chanjuan Gao, Yaqian Zhao......................................

1381

Adsorption of phosphate from aqueous solutions by thermally modified palygorskite Jingjing Xie, Tianhuhu Chen, Chengsong Qing, Dong Chen, Chengzhu Zhu, Jiayuan Wang, Xinmin Zhan …………………………………………………………………………..

1393

Treatment of an alkaline butyl rubber wastewater by the process of coagulation and flocculation -hydrolysis acidification - biological contact oxidation - MBR Yan Zhang, Wei Zheng, Rui Liu, Wei Li, Ying Li, Lujun Chen ……………………………………

1401

Structural evolution of heat-treated colloidal pyrite under inert atmosphere and its application for the removal of Cu(II) ion from wastewater Tianhu Chen, Yan Yang, Dong Chen, Ping Li, Yadan Shi, Xiao Zhu……………………………..

1411

Functionalization and immobilization of whole cell bioreporters for the detection of environmental contamination Cheng Chen, Dayi Zhang, Steven F. Thornton, Musen Duan, Yanjun Luo Aizhong Ding, Wei E. Huang......................................................................................................

1417

Composite vegetable degradation and electricity generation in microbial fuel cell with ultrasonic pretreatment Kun Tao, Xiangchun Quan, Yanping Quan ................................................................................

1423

Effects of COD:sulfate ratio on sulfate removal from oil shale retort water using microbial fuel cells Ho Il Park, Lian-Shin Lin …………………………………............................................................

1429

Occurrence and risk assessment of estrogens and anti-inflammatories in Baiyangdian Lake, North China Jianghong Shi, Xiaowei Liu, Jinling Cao, Ting Bo, Yingxia Li..................................................

1437

Shallow groundwater hydro-chemical evolution and simulation with special focus on Guanzhong Basin, China Wenke Wang, Lei Duan, Xiaoting Yang, Hua Tian.....................................................................

1447

Probabilistic scenario development to estimate future runoff in the Yellow River Basin, China Congli Dong, Gerrit Schoups, Nick van de Giesen ………………………………….....................

1457

Application of GIS in regional ecological risk assessment of water resources Yin Ge Liu, Ning Lian Wang, Lin Gang Wang, Ya Qian Zhao, Xiao BoWu …………………….

1465

Geochemical characteristics of reduced inorganic sulfur in a coastal environment, Bohai Bay, China Yanqing Sheng, Qiyao Sun, Wenjing Shi, Simon Bottrell, Robert Mortimer..............................

1475

Assessment of water environmental carrying capacity in Xi’an, China Jun Zhang, Ya-ni Geng, Qi Zhou, Ya-qian Zhao …………………………………………………....

1481

Solving water resources allocation problems using heuristic-based methods Yanan Jiang, Adrian T. McDonald, Martin Clarke, Linda See…………………………………….

1487

Ecological impacts induced by groundwater and their thresholds in the arid areas in Northwest China Wenke Wang, Zeyuan Yang, Jinling Kong, Donghui Cheng, Lei Duan, Zhoufeng Wang………………………………………………………………………………

1497

Estimating and modelling the sludge excess discharge in wastewater treatment plants in China Tao Xie, Chengwen Wang……………………………………………………………………………..

1509

Characteristics and feasibility study of sewage sludge for landscaping application in Xi’an, China Sheping Wang, Xinan Liu, Qin Zheng, Zhengliang Yang, Rixia Zhang, Bohan Yin……………

1515

Book Reviews Encyclopedia of Environmetrics (vol. 5) Abdel H. El-Shaarawi, Walter W. Piegorsch (Eds.)...................................................................

1521

Encyclopedia of Environmetrics (vol. 6) Abdel H. El-Shaarawi, Walter W. Piegorsch (Eds.)...................................................................

1525

ICEEM 07

ICEEM / 07

18 - 21 September 2013 Vienna, Austria

Location

For details see the conference site: www.iceem.eu, or contact the Conference Organizing Committee: [email protected]

Contacts

The fees for the ICEEM 07 Conference will include the conference materials (Program, Abstracts book), coffee breaks, lunches and the conference dinner. ŸFull fee: 300 EUR, ŸM.Sc. and PhD students fee: 200 EUR. This reduced fee will be available for 20 participants through competition. More details on www.iceem.eu

Fees

The extended abstract submission should be done no later st than 1 of March, 2013, via www.iceem.eu. After the review process, the accepted abstracts (2 pages) will be published electronically in a Book of Abstracts. The 2nd conference call and decisions on abstract acceptance st will be available after May 1 2013. Selected papers by the International Scientific Committee will be published in special issues of the journals: l Environmental Engineering and Management Journal (ISI impact factor 1.004) l New Biotechnology (ISI impact factor 2.867) l Energy, Sustainability and Society lInternational Journal of Nonlinear Sciences and Numerical Simulations (ISI impact factor 1.484) The Conference language is English.

Abstract Submission

The ICEEM 07 Conference will be hosted by the Vienna University of Technology in the beautiful city center of Vienna.

Faculty of Chemical Engineerng and Environmental Protection Department of Environmental Engineering and Management

(OAIMDD), Iaşi, Romania

ŸInterMEDIU Information, Consultancy and DL Department, TUIASI, Romania ŸAcademic Organization for Environmental Engineering and Sustainable Development

Biotechnology section

ŸEuropean Federation of Biotechnology, Environmental

Partners:

Vienna, Austria www.iceem.eu

18 - 21 September 2013

ICEEM / 07

1st Call for Papers

Integration Challenges for Sustainability

7 INTERNATIONAL CONFERENCE ON ENVIRONMENTAL ENGINEERING AND MANAGEMENT

TH

Faculty of Technical Chemistry Institute of Chemical Engineering Thermal Process Engineering and Simulation Research Division

Events

Prof.dr.,dr.h.c. Anton Friedl Vienna University of Technology

Environmental pollution and monitoring II. Water supply and wastewater treatment III. Environmental friendly materials IV. Sustainable processes and production V. Modeling, simulation and Plenary Sessions optimization Seminars VI. Environmental biotechnology Workshops VII. Waste management for resources and Posters Sessions energy recovery Social Events VIII.Environmental integrated management and policy issues.

I.

Topics

Prof.dr.,dr.h.c. Carmen Teodosiu “Gheorghe Asachi” Technical University of Iasi

Conference chairpersons:

The 7th edition of ICEEM continues its international and collaborative character, this time being organized by the “Gheorghe Asachi” Technical University of Iasi, Romania and the Vienna University of Technology, Austria, in close cooperation with the Environmental Biotechnology section of the European Federation of Biotechnology (EFB). This ICEEM edition seeks to bring together researchers, practitioners and specialists in various environmental fields, with a strong focus on integrating cutting-edge environmental technologies with efficient management practices, contributing thus to a sustainable future. Like before, ICEEM 07 strongly encourages contributions that focus on innovation, multidisciplinarity and cross-sectorial approaches related to environmental issues. Furthermore, the ICEEM conference welcomes contributions of young and senior scientists, and at the same time it encourages practitioners and specialists in various environmental fields to add a more practical-oriented approach to the conference sessions. We hope that the seminars, workshops, plenary sessions and sideevents of ICEEM will enhance multidisciplinarity, international cooperation and effective communication of scientists, engineers and managers.

Foreword

Dr. Catalin Balan Dr. George Barjoveanu Dr. Adela Buburuzan Dr. Daniela Căilean Dr. Petronela Cozma Dipl.-Ing. Adela Drljo

Organizing Committee

Prof.dr. Spyros Agathos Université Catholique de Louvain, Belgium Prof.dr. Thomas Amon Leibniz-Institut für Agrartechnik PotsdamBornim, Germany Prof.dr. Adisa Azapagic The University of Manchester, UK Prof.dr. Hans Bressers University of Twente, The Netherlands Prof.dr. Han Brezet Delft University of Technology, The Netherlands Prof.dr. Dan Cascaval Technical University of Iasi, Romania Prof.dr., dr.h.c. Yusuf Chisti Massey University, New Zealand Prof.dr. Cristina Costache Politehnica University, Bucharest, Romania Assoc.prof.dr. Igor Cretescu Technical University of Iasi, Romania Prof.dr. Göksel Demirer Middle East Technical University, Turkey Prof.dr., dr.h.c. Gheorghe Duca Academy of Sciences, Republic of Moldova Prof.dr. Emil Dumitriu Technical University of Iasi, Romania Prof.dr. Anca Duta Transilvania University of Brasov, Romania Prof.dr. Fabio Fava Alma Mater Studiorum-University of Bologna, Bologna, Italy Prof.dr. Maria Gavrilescu Technical University of Iasi, Romania Prof.dr. Ion Giurma Technical University of Iasi, Romania

Scientific Committee

Dr. Simona-Andreea Ene Dr. Michael Fuchs Dr. Raluca Hlihor Dr. Brindusa Robu Dipl.-Ing. Antonia Rom Dr. Walter Wukovits

Prof.dr. Christoph Herwig Vienna University of Technology, Austria Prof.dr., dr.h.c. Arjen Hoekstra University of Twente, The Netherlands Prof.dr., dr.h.c. Michael Jørgensen Aalborg University, Denmark Prof.dr. Emmanuel G. Koukios National Technical University of Athens, Greece Prof.dr., dr.h.c. Thomas Lindhqvist IIIEE, Lund University, Sweden Prof.dr. Matei Macoveanu Technical University of Iasi, Romania Assoc.prof.dr. Florica Manea Politehnica University Timisoara, Romania Prof.dr. Antonio Marzocchella University Federico II, Naples, Italy Prof.dr. Michael Narodoslawsky Graz University of Technology, Austria Prof.dr. Mircea Nicoara Al.I.Cuza University Iasi, Romania Prof.dr. Alexandru Ozunu Babeş Bolyai University, Cluj , Romania Prof.dr., dr.h.c. Ákos Rédey Pannonia University, Veszprem, Hungary Prof.dr., dr.h.c. Maria Madalena dos Santos Alves University of Minho, Portugal Prof.dr. Wilhelm Schabel Karlsruhe Institute of Technology, Germany Acad.prof.dr. Bogdan Simionescu Technical University of Iasi, Romania Romanian Academy, Romania Prof.dr. Krzysztof Urbaniec Warsaw University of Technology, Poland


July 2013, Vol.12, No. 7, 1309-1310



EDITORIAL A SPECIAL ISSUE ON RECENT ADVANCES IN WATER RESOURCE MANAGEMENT AND POLLUTION CONTROL: WITH SPECIAL FOCUS ON CHINA A fast economic development in China over the last three decades has lifted hundreds of millions of people out of poverty. However, an annual GDP increase of over 10% due to the large scale industrial activities and fast urbanization has released substantial amounts of pollutants into the environment and exerted tremendous pressure on local ecosystems. Although a large amount of investment has been made by the Chinese government and a series of efforts have been made by the environmental regulators, scientists and engineers, the current situation in China is still challenging. The water in many major rivers cannot be used as drinking water; More than half of the major lakes cannot be considered as drinking water sources as well due to high levels of nutrients (N & P); Many water bodies are even unfit for irrigation; Soils in key agricultural regions are contaminated by heavy metals; Scarce arable land and water resources and important biodiversity are being lost; China’s carbon and nitrous oxide emissions are rising; The air quality in many major cities is worrying. China urgently needs to shift to a more environmentally sustainable model. Full efforts should be made to consider the future growth without further undermining ecosystems and the natural resources. Among them, water resources management and water pollution control is critical in China and will be of increasing importance in the future. This special issue on recent advances in water resource management & pollution control in Environmental Engineering and Management Journal (EEMJ) provides a good opportunity to showcase the most recent research and development on China’s water

resource management and pollution control by Chinese scholars (mainland and overseas) and their colleagues worldwide. The issue begins with a holistic view on China’s water management – challenges and solutions. It is followed by different sized research on wastewater and sludge treatment, including novel process development, deep understanding of the treatment mechanisms and cost-effective treatment optimization. Thereafter, the issue presents a number of studies on China’s water resource management in regional and national levels. The authors are invited to contribute their most recent research findings and practical technologies which can most likely be applied in China. We hope that water resource management and pollution control in China will gain a new momentum, if any, through this special issue. We would like to thank all the authors invited for their participation, this enriching the diversity of perspectives and contents of this special issue. In particular, we highly appreciate the great support from Professor Matei Macoveanu, Editor-in-Chief of EEMJ, and Professor Maria Gavrilescu, Executive Editor, for providing the valuable journal volume to dedicate a special issue. Thanks are also given to a number of reviewers. Their precious time and invaluable and detailed suggestions have been especially helpful in improving the quality of each paper and therefore this special issue. In addition, we wish to sincerely thank the Irish EPA, Research office and Ryan Institute at NUI, Galway, for their financial support towards the publication of this special issue.

Editorial/Environmental Engineering and Management Journal 12 (2013), 7, 1309-1310

Gathering of some contributors Guest editors: Dr. Yaqian Zhao University College Dublin, Ireland Dr. Jidong Gu The University of Hong Kong, P.R. China Dr. Xinmin Zhan National University of Ireland, Galway Dr. Yaqian Zhao is the Director of Water and Effluent Laboratory in the School of Civil, Structure & Environmental Engineering, University College Dublin (UCD), Ireland. He received BEng. (1984) and MEngSc. (1990) degrees in China and PhD in Strathclyde University in Scotland (2000). He then worked as Postdoc research fellow in Queen’s University of Belfast (N. Ireland, UK) from 2000 to 2004 before joining UCD in 2004. His research covers a number of issues in broad area of water, wastewater and biosolids/residual treatment engineering with specific achievements in (1) constructed wetland systems for wastewater treatment; (2) phosphorus removal/ immobilization/ adsorption and (3) waterworks sludge conditioning, dewatering and beneficial disposal/reuse. Dr. Zhao is an active researcher in international level with over 170 research papers published in refereed journals (highest IF 7.409), book contributions and international conferences. He has been invited to deliver near 80 lectures and research seminars internationally (including Cambridge University, UK). Dr. Zhao is a journal founding member and three journal editorial board members as well as 37 journal reviewers. Dr. Ji-Dong Gu is Associate Professor in the School of Biological Sciences, Faculty of Science, The University of Hong Kong, P.R. China. He obtained his BSc degree in PR China (1983), MSc degree from University of Alberta in Canada (1988) and PhD degree from Virginia Tech in USA (1991). He then worked in the Microbial Ecology Laboratory of Harvard University between 1993 and 1998 before joining The University of Hong Kong in 1999 as Assistant Professor and then in 2004 promoted to Associate Professor. He has published in the more than 190 scientific journal refereed papers, 32 book chapters, co-edited a book with Ralph Mitchell on Environmental Microbiology (2nd ed, John Wiley-Blackwell. 2010), wrote a book on Biosusceptibility of Polymers and Fiberreinforced Composites and Testing Methods (Springer, 2014) and edited/co-edited special issues in International Biodeterioration & Biodegradation and Ecotoxicology. Dr. Xinmin Zhan is a lecturer (above the bar) in the Environmental Engineering Group, NUI Galway. He studied Environmental Engineering at Tsinghua University, Beijing, China for his BEng and PhD degrees. His research interest includes: (1) Development of cost-effective and efficient technologies for domestic and industrial wastewater treatment; (2) Recovery of organic waste and biomass for use as biofuels; and (3) Interactions between climate change and the waste treatment infrastructure. He has published over 65 journal papers and has given over 80 invited lectures, conference presentations, and seminars.

1310


July 2013, Vol.12, No. 7, 1311-1321



CHINA’S WATER MANAGEMENT – CHALLENGES AND SOLUTIONS Yanan Jiang1,2,3, Faith Ka Shun Chan2,3,4, Joseph Holden2,3, Yaqian Zhao5, Dabo Guan3,6 1

College of Water Resources and Architectural Engineering, Northwest A&F University, Yangling 712100, P.R. China 2 School of Geography, University of Leeds, Leeds LS2 9JT, UK 3 Water@Leeds, University of Leeds, Leeds LS2 9JT, UK 4 The University of Nottingham Ningbo China, 199 Taikang East Road, Ningbo, P.R. China 5 UCD Dooge Centre for Water Resources Research, School of Civil, Structural and Environmental Engineering, University College Dublin, Ireland 6 School of Earth and Environment, University of Leeds, Leeds LS2 9JT, UK

Abstract China has experienced enormously rapid development since the open door policy introduced in 1979. Population has increased by 30% to 1.3 billion, and the annual GDP growth rate was 9.8 % in the last few years. However, frequent water disasters in recent years have caused significant damages to China’s regional growth and societies. There are huge contemporary challenges for Chinese water resource management. In this paper, we examine three major challenges for China’s water resource management, which are water scarcity, water pollution and flood management. We discuss some of China’s past management strategies and its future water management plans which come with major new investment. China will be further developed during this century and we provide some thoughts on water resource management that could be undertaken in China to increase resilience in face of a capricious future. Key words: China, climate extremes, flood risk, pollution, water resources management, water scarcity Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Management of China’s water resources is highly challenging because of rapid population growth (over 30% since 1979), rapid urbanization and large areas of dense population, enhanced expectations of standards of living, economic growth, eastward migration, and a highly variable climate which is difficult to forecast (Chong et al., 2012; Hubacek et al., 2009; Varis and Vakkilainen, 2001). The result is a situation of: i) water scarcity in many areas which also impacts food security; ii) water body deterioration/degradation; and iii) vulnerability to climate extremes including droughts and floods, and climate change impacts including global sealevel rise and storm surges which can have major



impacts on the mega-coastal cities of China (Nicholls, 1995; Varis et al., 2006). China’s economy has been growing rapidly since the “open door policy” which started in 1979 and the pace of growth increased after the industrial expansion from the mid-1980s (Chen et al., 2011; Feng et al., 2012; Guan et al., 2008). Within around thirty years, China developed from a country with heavy poverty and severe production inefficiencies to become the 2nd largest economy in the world (Feng et al., 2009). Inevitably, China has paid an enormous price for its success in terms of environmental degradation and natural resource depletion (Guan and Hubacek, 2007; Yu, 2010). Along with all its impressive achievements, China is facing severe resource challenges, especially in water management.

Author to whom all correspondence should be addressed: E-mail: [email protected]

Jiang et al./Environmental Engineering and Management Journal 12 (2013), 7, 1311-1321

Economic reforms have transformed the nation to enhance rapid urbanization and economic development (Minx et al., 2012). For example, it created job opportunities to attract large migration from midele and west to the east and south. Unfortunately, this rapid socio-economic movement also creates huge challenges for water management, such as freshwater consumption from industrial and domestic users (Guan and Hubacek, 2008). Many cities, particularly in coastal and deltaic areas, are over-extracting groundwater resources to fulfill their demands, resulting in severe landsubsidence. Subsidence then acts as a catalyst for increasing flood vulnerability by storms (e.g. typhoons), surges and sea-level rise (Fuchs et al., 2011; Syvitski, 2008). Groundwater abstraction and subsidence in coastal areas and in China’s vast and populous deltas also results in salt water intrusion which has resulted in severe water shortages in some years such as the Pearl River Delta problems in the recent years from 2006 (Chen et al., 2009; Luo et al., 2007; Xu et al., 2009). This phenomenon demonstrates the interconnectedness of current water problems in China. There is therefore an urgent need for holistic water management strategies that tackle the interconnected nature of water problems. The Chinese government has been trying to solve water scarcity and uneven water distributions between different geographical regions in the country. There are some large scale hydro-engineering projects such as the South-North Water Transition Project to relieve huge water demand and consumption in the North. This project will deliver more than 1.76 million m3, benefiting more than 325 million people in the North China Plain (Khan et al., 2009). However, such water transfer projects also create social and environmental problems including population resettlement and adverse environmental impacts which present huge challenges yet to be tackled (Jiang, 2009). In this paper, we present the three most important water management challenges in China, and traditional policy barriers to progress. Some possible solutions provided by China’s new water policies will be discussed and evaluated in terms of alignment to tackling integrated water resource challenges. 2. Three water management challenges China’s sustainable development strategy report (Chinese Academy of Sciences, 2007) indicated that 63% (420) of China’s 667 cities have water shortages (water scarcity), 30% (200) have extremely seriour water shortage problems, this number was one sixth (110) back to 2003. According to the latest Water Resources Bulletin (MWR, 2012) 35.8% of all the rivers are severely polluted and 71 of all the 103 lakes suffer from eutrophication (nutrient enrichment leading to blooms of aquatic plants at the expense of other organisms). Water related disasters such as the 1998 floods and 2010 droughts have

1312

caused large losses of life and property, as well as damage to human wellbeing and the environment. Water scarcity, water pollution and flooding are the three major challenges for Chinese water management. These three challenges have been widely discussed (Jiang, 2009; Khan and Lui, 2006; Lee, 2006; Liu and Raven, 2010; Liu and Yang, 2012). There are nearly 40 reports available from the World Bank commenting on a range of China’s water issues such as water pollution, water pricing, ecological compensation, water rights, water governance and strategy. The challenges of water scarcity, water pollution and flooding are described below. 2.1. Water scarcity There are three indicators of quantity related water scarcity in China: water shortages, water resources overexploitation, and the effect of water resource overexploitation on the environment. China’s water shortages date back to the 1980s since the fast economic development and rapid urbanization strated and different scales and degrees of water shortage problems have been emerging and increasing until now (World Bank, 2007). As the population is still increasing and the urbanization process is still going very rapidly, the projected water shortage problems will be even worse if this problem was not taken seriously. Nearly all (30 of 32) metropolitan citys (more than 1 million population) have difficulties in meeting their water demands (Li, 2006). China's total water deficit could reach at 40-50 billion m3 by 2030. The water shortage in the 3H basin areas (Huang River – Yellow River, Huai River, Hai River) is projected to be 56.5 billion m3 by 2050. Annually water shortages in the 420 water scarcity cities alone have caused 200 billion RMB economic loss. Scarce water has been overexploited in North China (Feng et al., 2012b). In 2010, North China obtained 81.1% of its water supply from surface water and 18.4% from groundwater. In terms of water development ratio (the ratio between water supply and water availability) these range from 25% to 120% (the high figure is for the Hai River in 2010) for basins in the north compared to rates of 18 to 66% in South China (MWR, 2012). Surface water overexploitation has reduced instream flows and caused negative impacts on the aquatic ecosystems. For example, 40% (about 4000 km) of the watercourses in the Hai River basin have dried up and 194 natural lakes disappeared (Wang, 2000). The annual average discharge (from rivers to the ocean) nationwide has substantially decreased from 24 billion m3 in the 1950s to 1 billion m3 in 2001 (Xia et al., 2007). River flow in the Yellow River has decreased 51% since the 1950s (Wang and Jin, 2006; Fan et al., 2006). In 1997, the lower reach of the Yellow River had no flow for more than 226 consecutive days; the length of the main channel with no flow was 700 km which accounts for 90% of the

China’s water management – challenges and solutions

river course in the lower reach (Feng et al., 2012b; Wang and Jin, 2006). As a result the Yellow River Delta has become more fragile and susceptible to natural hazards (Deng and Jin, 2000; Fan et al., 2006; Huang and Fan, 2004; Lin, 2001). Groundwater has been severely over abstracted. Since 1980, regions with over extraction of groundwater have increased from 56 to 164, with the total area increased from 87,000 km2 to 180,000 km2 (MWR, 2007b). A percent of 70 (or 90,000 km2) of the North China Plain has been affected by over extracting groundwater (Liu and Yu, 2001). As a result, there have been two additional negative environmental effects: seawater intrusion and ground subsidence. Falling groundwater tables in coastal regions could break the balance between freshwater and seawater and cause seawater intrusion problems. According to the State Oceanic Administration People’s Republic of China (SOA, 2013) seawater intrusions has occurred in 9 coastal provinces (Liaoning, Hebei, Tianjin, Shandong, Jiangsu, Shanghai, Zhejiang, Hainan, Guangxi) with the longest instrusion distance 32.1 km in Shouguang city Shandong province, covering a total area of 2000 km2. Ground subsidence is another problem caused by groundwater over abstraction. According to Shalizi (2006) ground subsidence was observed in northern and eastern China. In major cities like Beijing, Tianjin, and Shanghai the subsidence coule be several meters. It is estimated that from 20072011 21 billion RMB has been invested in Shanghai to prevent or overcome this problem. The drop in water table is also associated in many old urban centres with oxidation of organic deposits which decays them and also leads to subsidence (An et al., 2012; Holden et al., 2006). In future, climate change may make the situation even worse as areas where water is scarce will become even drier. In the Yellow River basin, average temperatures have increased while precipitation and river runoff have decreased over the past 50 years (Fu et al., 2004; Xia et al., 2004; Yang et al., 2004). In the past 20 years, climate change has decreased water resources in northern China. In addition, the loss of glaciers and wetlands from the Qinghai-Tibetan Plateau has decreased river runoff by 917 billion m3 in total over the past 50 years and will lead to an annual loss of 143 billion m3 in the future (An et al., 2006). At the same time, China’s water demand may increase by 6.5%, 32%, and 35% (2003-2020) from agriculture, industry, and residential users, respectively. By year of 2050 even considering the necessary water saving measures such as improving water use efficiency in agriculture and industry, the total water demand will increase from 570.2 to 832.3 billion m3, with the industrial water demand may still increase from 66.5 to 343.6 billion m3, while the agricultural water demand hopefully will decrease from 484.8 to 415.7 billion m3. At the same time residential water demand will increase from 18.9 to 73 billion m3 (Liu and He, 2000). So it can be

predicted that in future China will face even more severe water scarcity challenges. 2.2. Water pollution Water pollution is another water challenge which needs to be addressed. China’s lakes and reservoirs have been experiencing accelerated eutrophication and degraded water quality (Jin et al., 2005). Chinese authorities classify water quality into five grades (from the best quality at Grade I to the worst quality at V) based on the purposes of use and protection targets. Based on this classification, water quality is being monitored on a regular basis in almost 500 monitored stations through national and provincial water monitoring centers. According to the 2011 Water Resources Bulletin (MWR, 2012), of all the 103 monitored lakes, 58.8% met the standards at good water quality (Grade I to III) and 41.2% were ranked poor (Grade IV & V). The three major lakes including Tai, Chao, and Dianchi are the most polluted lakes in the country with water quality even below Grade V. As for the river water quality only 64.2% of monitored river water was in categories I to III, while as much as 22.9% was in the worst two categories (MWR, 2012). More than 40% of China’s rivers are severely polluted and more than 50% of its lakes suffer from eutrophication (Liu and Yang, 2012). There are several severe consequences of water pollution. It causes water quality related scarcity. This kind of water scarcity has occurred in northern and eastern China. Shanghai, located downstream on the Yangtze River and the Lake Tai basin, has its water polluted from upstream and the local area around the lake (Wang et al., 2010). Zhejiang Province faces the same problem: water scarcity here is not because of a lack of water to use, but poor quality renders water unusable. Qu and Fan (2010) noticed the river pollution in urban rivers are particularly difficult to control, now up to 80% of urban rivers in China have been contaminated to varying degrees by pollutants such as nitrogen, phosphorous, organic compounds, and heavy metals. According to World Bank (2007), 25 billion m3 of water was not used because of poor water quality from 2000 to 2003. 10% (47 billion m3) of China’s total water supply (563.3 billion m3) in 2005 came from degraded supplies, which means they can not meet the before-treatment quality. Degraded water quality has caused serious impacts on both society and economic. In 2003 poor water quality alone has caused at least 158 billion RMB (World Bank, 2007). 2.3. Flood management Floods in the large river basins, coastal lowlying areas and megacities (population commonly more than 8 million) are major problems. China has a long history of facing floods. Jun and Chen (2001) reported the country suffered large scale severe flooding more than 1000 times from 206 BC to 1949

1313


AD. Zhang (1999) reported more than 2000 lives lost from 1990 to 1999 and the monetary figure was over 800 billion RMB (1 USD roughly equals to 6.4 RMB) from flood disasters, equivalent to 20% of total revenue for that period. There are two main reasons why China suffers in the face of big floods: (i) natural factors (i.e. climate, topography, etc.) and (ii) human induced factors (i.e. populations and urbanization). 2.3.1 Natural drivers of flooding The natural factors to influence flooding in China include the monsoon climate from the Pacific Ocean, topography and the geomorphology (e.g. sediment flux and drainage capacity) of major rivers (i.e. Yangtze River, Yellow River and Pearl River). For example, more than 1800 mm with about 80% of annual rainfall falls between May to September during the wet season in the Pearl River Delta (PRD). The Yangtze River Delta also records more than 1100 mm in the summer season (Gu et al., 2011), while some heavy rainstorms could produce over 400 mm rainfall in 24 hours (Lee et al., 2010). These climatic conditions cause serious urban floods if cities do not have well-equipped stormwater drainage systems. The Beijing urban flood event on 21 July 2012, with more than 300 mm of rainfall in one day, caused 37 deaths with pluvial flood damage claimed to be because of a lack of an up-to-date urban drainage capacity (Nan, 2012). Large scale fluvial floods in the Yangtze River Basin during monsoon periods have resulted in the inundation of cities like Wuhan (Jun and Chen, 2001). The 1998 floods saw 29 provinces flooded, 21.2 million ha land inundated and an astonishing total of 223 million people losing their homes. The direct economic loss was more than 166.6 million RMB. The indirect damage to wider economies is not measurable at moment. The coastline of China is about 18,000 km (Fuchs et al., 2011) and many megacities like Guangzhou, Hong Kong, Shenzhen in the PRD, Shanghai in the YRD, Qingdao nearby the Yellow River Delta and Tianjin occur in coastal areas. Recent research predicts that most of these Chinese coastal cities will suffer from enhanced flood risk to massive populations and economic assets over the coming 60 years (Nicholls et al., 2008). Coastal flooding usually results from typhoons and storm surges. Recorded storm surges can cause tides higher than 1 to 3 metres above the normal high tide level in Guangdong Province (Zhang et al., 2011). More than 3720 km2 low-lying flood prone areas in the PRD are currently vulnerable to coastal inundation (Syvitski et al., 2009). Low-lying coastal areas in Hong Kong such as Tai O town and Shengwan business district were inundated three times by storm surges between 2006 and 2012 which caused substantial economic loss. Lu and Yao (2006) reported Typhoon YunNa in 2004 caused storm surges more than 3 metres above the peak tide level along the Zhejiang coast in YRD, sea walls were breached in more than 1200 locations damaging more than 42,300 ha of farmlands, 3,000

1314

fishing boats and 200 docks; more than 467,000 people were evacuated. Global sea-level rise and frequency of surges are predicted to create further impacts to coastal areas in China. Sea-level is predicted to rise between 200 – 300mm in the Guangdong and PRD region by 2030 (Zhang, 2009), 380mm in Shanghai area and about 340mm in Jiangsu Province (Gu et al., 2011). Salt water intrusion is also likely to increase as a result of sea-level rise (Mimura and Kawaguchi, 2011), contaminating groundwater and freshwater supplies, and particularly threatening coastal cities which rely highly on groundwater resources such as Shanghai and Tianjin. 2.3.2. The human-induced drivers of flooding Since the economic boom started in the 1980s, most of the south and east coast deltaic regions in China have been developed as Special Economic Zones (SEZs) (Yeung, 2011) to provide special tax rates for foreign companies and attract foreign investment. The SEZs such as Shenzhen have good logistical advantages (e.g. international ports, railway terminals and airports) aiding international cooperation (Chan et al., 2012). These places provide good employment and education opportunities to attract a large amount of human resources, enhancing internal migration towards the south and east coast in China (Bailey, 2010). Taking the Pearl River Delta as an example, we note that the population in 1979 with its 11 cities including Hong Kong and Macau was not greater than 15 million, while now it is about 60 million (Yeung, 2010). Shenzhen was a fishing town before 1979 with about 100,000 people. Shenzhen is now a megacity with a population of about 8 million, alongside nearby Guangzhou (12 million) and Hong Kong (7.5 million) (Ng et al., 2011). All of the rest of PRD’s cities are currently with at least 3 million population (Yeung, 2010), and the delta is transforming to be the most populous delta in East Asia with a population of more than 120 million by 2050 (UN-HABITAT, 2008). There is a similar pattern in the YRD area with more than 97 million people. Shanghai (15.7 million) is the most important economic hub in the YRD region and more than 24 cities in the YRD have more than 1 million people. The region also accounts for 17.5% of the whole country’s GDP with 4.3 trillion RMB in 2008; the GDP per capita was more than 44,468 RMB (2 times higher than the national average level) (Ge et al., 2011). Gu and Han (2010) also reported on the other large coastal economic zone – Bohai Economic Rim, with megacities such as Beijing (18 million), Tianjin (10 million), Qingdao (7.3 million) along with another 8 cities with at least 1 to 6 million people (e.g. Jinan, Dalian, Shenyang, etc). The Rim, with more than 75 million people, contributes over 4710 billion RMB GDP with more than 62,000 RMB GDP per capita. All three mega economic zones in China are closely located in coastal flood prone areas. While these areas are


naturally flooded frequently, the rate of population and economic growth, plus future climate change impacts, means that the new flood risk in these areas will be a massive challenge to manage (Fuchs et al., 2011; McGranahan et al., 2007). Seto (2011) cautioned that current practice and policy may not be enough to cope with such fast urbanization rates in Asian mega-deltas including the YRD, PRD and Bohai Economic Rim. Chan et al. (2012) found about 7 Chinese coastal cities (Guangzhou, Shanghai, Tianjin, Ningbo, Shenzhen, Qingdao and Hong Kong) are ranked at the top 20 global port cities that exposed to coastal flood impacts with population and economic assets. The Research Team of China Climate Change Country Study (1999) estimated that a sea-level rise of 300mm accompanied by a 1-in-100 years storm surge would cost more than 231 billion RMB economic losses and inundate more than 80% of the PRD area (6,520 km2) (Zhang, 2009). Like other coastal areas in China, the PRD is highly reliant on hard engineering infrastructure (e.g. dikes and embankments) for flood protection. Cai et al. (2011) doubted that most of the protection measures in the region would cope with a 1-in-100 years flood return period, particularly with sea-level rise over the next 60 years. Rapid urbanization in coastal areas enlarges flood risk because flood water storage capacity is reduced by massive land-use change and land subsidence. The PRD, for example, has increased urbanized land from 29.5% in 1982 to 80% in 2010 (Chan et al., 2012). Recent reclamation project along the PRD coastline (i.e. Shenzhen Bay area) has reflected land scarcity in the region. However, the authority has removed some parts of the wetland (i.e. mangroves) area to reclaim (Zhou and Cai, 2010), which may further reduce water storage from tidal changes and resulting higher vulnerability to the new developed infrastructures. Shi et al. (2007) reported such urbanization has largely declined the soil infiltration, water storage capacity and increased the surface runoff. As a result the maximum flood discharge has increased nearly 12.9% on average over the past two decades. Such development rates also enhance long-term over extraction of groundwater, especially in the coastal megacities like Tianjin and Shanghai which lack freshwater resources (Hu et al., 2004). The groundwater levels in these areas have been substantially decreased with associated land subsidence. Tang et al. (2008) reported the coastal area in Shanghai has subsidence rates in some places over 2 m since the 1960s. Most cities in the YRD such as Suzhou, Wuxi, Changzhou, Nantong, Hangzhou and Nanjing have all subsided by 0.5 to 1m. 3. Traditional water resource management The spatial and temporal distribution of water resources is not consistent with socio-economic water needs which cause conflict between water

supply and demand. Poor water resource management makes the situation even worse. In China water was managed by multiple government agencies at different authoritative levels. Lack of effective coordination and cooperation among them led to a fragmented system which could not manage water resources effectively. China's water resources administration was divided between the State Environmental Protection Administration (SEPA) and the Ministry of Water Resources (MWR). SEPA was responsible for controlling water pollution, while the MWR was responsible for water resources planning, including designating water functional zones for different uses and establishing corresponding water quality standards. The coordination between them was severely inadequate which means water quantity and water quality was managed by different authorities which impeded efficient water resource management. Integrated water resources management based on river basins has been commonly accepted as an effective approach for managing water resources. In China although commissions for major rivers and lakes were established to promote integrated management, they had limited power to allocate water resources, coordinate water resource exploitation and conservation, and enforce water resource planning at the basin level. The authority and responsibilities among these government agencies was not clear enough, and this undermined their ability to regulate water resources management, which led directly to a water resources management largely based on political boundaries rather than on watersheds, which amplifies these issues. China’s water resources management is still driven by water supplies which ignores the economic nature of water resources. With economic development and population growth, this passive management with no restrictions on water demand has led to inefficient water use. For example, in 2003, China's water use per 10,000 GDP was 4.5 times that of most developed countries. China's average recycling rate of industrial water use was estimated to be 40-50% which is low compared to 80% in developed countries (CAS, 2007). In agriculture, as indicated by CAS (2007) and Zhang et al. (2007), the ratio of actual irrigation water consumption to the amount diverted is only 0.45 which is far below the level of 0.7 to 0.8 in developed countries. China does not have a water rights system which can promote effective water resource management. Jiang (2009) attributed much of the water use inefficiency and the current water scarcity in China to this issue. Back to 2000 strengthening water rights development has started such as revising the water law and issuing policy guidance (FAO, 2001), however until now water rights are still incomplete by modern standards. Efficient and effective systems have not been established to manage the three components of water rights: the amounts that can be withdrawn, transferred, and must

1315


be returned with certain quality. Initial water rights based on water allocation have not been completed, rules and methods regards to sustainable water allocation are still in complete and legal delineation about water withdrawal permits is unclear. Not all water uses are measured and managed by permits. Water withdrawal permits are not consistent with water allocation due to lack of coordination mechanism exists within basins. Water rights trading has taken place and efficient water resources allocation was improving, however water rights trading was still largely based on administrative commands instead of water markets, as a result the scale of water rights trading was relatively small and inefficient. Relatively low water price does not place incentives to end water users to conserve water. China’s water prices were set through a political administration instead of by market, so the price is relatively low. Water prices were purposely set low and are insufficient to cover the full cost of water supply. It is estimated that current household expenditures for water only account for about 1.2% of disposable income. This percentage is lower than the 2% level that stimulates water-saving behavior and is much lower than the 4% in developed countries (Zhang et al., 2007). These low water prices provide little or no incentives to save water. In addition to the issues mentioned above, policies are not well integrated with each other and may exacerbate water resource issues. Many policies, including urban planning, industrial development policy, agricultural policy, etc., can have indirect effects on water resources. If these potential effects are not accounted for, policy outcomes will likely be inconsistent with the carrying capacities of local water systems. The present discrepancy in the distribution of socio-economic development and water resources is a typical example of policy failure to consider water resources. 4. Solutions to water challenges 4.1. Solutions to water scarcity and water pollution The characteristics of China’s water resources determine that huge hydraulic engineering programmes need to be implemented to maintain China’s socio-economic development. Therefore the first solution is to change the natural water availability via engineering measures. There have been quite a lot of achievements: (1) by building reservoirs and dykes China now has a flood control system for key rivers; (2) the total irrigated land reached 60 million ha which supports national food security; (3) China now has 660 billion m3 total water supply capacity which provides water for industrialization, urbanization and socio-economic development; (4) China now has very large sacle of water transfer projects under construction. To improve water resources management through engineering is a cost-effective option that

1316

can alleviate China’s water scarcity. However, there are several suggestions with regards to this option. First, a clearly defined legally enforceable water rights system need to be developed which can make the regulation of water withdrawal and use effectively and efficiently, an institutional system is recommended. Second, a water market where the water rights can be traded is necessary to improve water use efficiency. Comparing expensive engineering measures this approcach is another quite cheap and potential way. Third, sustainable water resource management need to play a very important role in a near future, so research-based, data-driven decision support systems, and information system should be paid more attention in future government plans. Integrated water resource management (IWRM) could be a useful approach. China has begun to adopt IWRM concepts. China has been working together with the European Union (EU) to enhance IWRM through EU-China river basin management programs through which China will share EU’s experience and best practices. IWRM can enhance the integration of flood management and environmental management objectives. Appraisals of environmental impacts during planning and design of future flood management schemes should be conducted to minimize adverse environmental impacts but also to enhance environmental (i.e. nature and biodiversity) and social (i.e. water supply, recreation in flood retention or wetlands) outcomes. Liu and Yang (2012) have proposed some policy recommendations: (1) Integrated monitoring and proactive measures: Indicators that directly and indirectly drive water quantity and quality changes should be monitored. Ministries should expand monitoring to indicators in human dimensions (e.g., values and attitudes toward water, land use, and development). (2) Integrating social sciences: China should integrate natural sciences and technology with social sciences which may help in predicting water demand and long-term effects of China’s water plan, and could lead to positive changes in human behaviors. (3) Enhancing international cooperation: China’s water plan should consider both virtual water and real water trade. 4.2. Flood management The most common practice in China - flood engineering protection - is expensive and unprecedented urbanization in flood prone areas has meant that Chinese authorities have been unable to keep pace with the risk. Therefore, a long term, more resilient, cost effective and sustainable flood risk management strategy is needed in Chinese megacities and nationwide. Flooding is a natural process but it can be exacerbated by land management. In China agencies often only deal with regional, city or town planning development issues rather than integrate catchmentbased flood issues. In the United Kingdom, where


this was also true, more than 8 million properties were built on the floodplains and frequently suffer from coastal and inland fluvial floods because urban development and town planning were misplaced in high flood risk areas (Pitt, 2008). The UK government has improved the practice with implementing flood risk appraisal practice, such as inclusion of Policy Planning Statement 25 (PPS25) to modify flood hazard and the possible exposure of flood risk in planning practice. This means the private developer or public authority is obliged to assess possible flood risk before submitting any new development proposals. The practice has also addressed the potential flood risk impact in the proposed development area for different stages (after 10, 25, 50 and 100 years) to ensure the development plans are able to be resilient, sustainable and adaptable to uncertain climatic changes (McLean and Watson, 2009). Therefore, in China, the Ministry of Water Resources (MWR) may possibly consider reform and merge the flood management section with the planning authority or establish working partnerships between two Bureaus to coordinate ongoing flood management planning and implement flood management into the land-use management policy. Soft flood protection measures such as flood forecasting, flood risk mapping and hazard zoning systems, multi approach decision support systems, flood warning systems, and flood emergency and excavation plans can effectively mitigate flood risk, reduce casualties and economic losses around the world (Plate, 2007; Meyer et al., 2009; Harvey et al., 2009; Kreibich et al., 2011; Gersonius et al., 2011). Cheng (2006) agreed that China should apply these soft measures. In fact, measures like flood forecasting have been implemented since the 1980s in China but the accuracy of the forecasting system could be improved with more precise data monitoring networks and up-to-date hydraulic modelling methodologies (i.e. two-dimensional hydraulic models in Shenzhen River) (Chan and Lee, 2010). The forecasting procedure can be connected with flood mapping to provide more detail about extents of inundation impacts. Such practices have been applied in the EU Floods Directive (Marchi et al., 2010). Broader application of flood risk mapping is required in China as this measure provides enormous support to non-structural flood protection measures which includes land-use management, community awareness and preparedness, and potential flood insurance schemes; as well as emergency excavation plans. Government officials and decision makers can then build a decision support system from these advanced and accurate flood risk maps and forecasting information. The public (community) can be alerted by flood risk mapping and forecasting systems to raise their awareness of flood risk and encourage their engagement with flood preparation and emergency excavation plans (White et al., 2010).

Ma et al. (2010) reported that non-structural flood protection measures including flood information collection, warning systems, planning of disaster prevention and emergency response or excavation plans have been enhanced in the mountainous parts of China to reduce impacts of flash floods, as it is difficult to cover such huge area of 4630,000 km2 for the 556 million populations by building flood protection infrastructure. The UK Institute of Civil Engineers (ICE) has suggested that it is impossible to avoid any kind of fluvial flood risk under uncertain climatic regimes from a flood engineering perspective. They suggested “living with rivers” to all governmental officials (Fleming, 2002) which equates to allowing rivers to flood, but adapting and becoming resilient to flooding. People may need to learn to accept some degree of flood risk in return for the benefits to be derived from using land subject to flood risk. Floods are periods of renewal for nature and are a natural land-forming process to sustain many aquatic ecosystems and benefit biodiversity. Therefore, flood engineers and planners should consider if alternative land could be used as flood retention and ensure that building developments are designed to cope with floods in flood prone areas (e.g. buildings with car parks on the ground floor which are easily evacuated during floods; but living space higher above likely flood levels). 5. China’s new water policy The Chinese government is aware of the challenges and has started to reform water resources management to address these issues. However due to complexity, the challenge is still severe. Within its 11th Five-Year Plan (2006-2010) a series of policy goals and priorities for water resource management was proposed under the direction of “scientific development” and “harmonious society” which is the general goals and guiding principles for the whole plan (State Council, 2006). Policy objectives for water resource management was established such as strengthening river basin management, improving water use efficiency in agriculture, protecting drinking water sources and increasing the urban sewage treatment rate etc (State Council, 2006). The 11th Five-Year Plan for Water Resources Development includes both action plans and methods for implementation (MWR, 2007b), such as expediting water allocation, developing water rights systems, implementing quota and demand-side management, and improving water use efficiency and benefits (Geng et al., 2011), more importantly it reflected a strategic shift towards sustainable water resource development. In January 2011, the government’s annual “No. 1 Document,” which reflects its top priorities, outlined a plan to expedite water conservancy development and reform and to achieve sustainable use and management of water resources within this

1317


decade. The plan includes total investment of four trillion Yuan (U.S. dollar 635 billion) to solve water problems in the next 10 years. This plan has huge significance for the whole nation and sets goals, policies and tasks for the next 10 years. This document states: (1) Water is the origin for life, the vital element for production and the foundation for ecology; (2) Water is of the characteristics of public interests, basic and strategic functions. In future the Chinese government will focus on the following issues: (1) highlight the safety and quality of people’s livelihood; (2) emphasize equality of public water services; (3) focus on improvement of vulnerable aspects in respect to water; (4) pay more attention to sustainable use of water; (5) implement strict water management; (6) construct a water-saving society; (7) protect water resources; (8) create compatibility between bearing capacity of water resources and the economy; (9) enhance capacity building for organizations; (10) promote participation of the public in water management. The No. 1 document of the 12th Five-Year Plan (2011-2015) (State Council, 2011) has resulted in some actions to achieve the overall goals outlined above. These include: (1) Establish a flood and drought control and relief system. (2) Establish a reasonable water resources allocation and highly efficient water-use system. (3) Establish a system to protect water resources and secure the health of aquatic ecosystems. (4) Establish a system and mechanism that facilitates the development of water conservancy. Detailed information has been translated which can be acquired via https://www.sciencemag.org/content/suppl/2012/08/0 8/337.6095.649.DC1/1219471.Liu.SM.pdf One of the additional ways forward is to integrate different components of the water challenges solutions. For example, turning floodwater and its pollutant load into a resource for water supply or fertilisers would be a potentially “win-win” strategy. Therefore the Chinese authorities seek to develop integrated infrastructure that serves multiple functions such as flood control, water storage and water treatment and re-use facilities. 6. Conclusions It worth menthion again that the Chinese government has decided to invest more than 4 trillion RMB (~ 600 billion US$) in infrastructures (including significant portions of finance to water related infrastructures) over the next decade; this demonstrates that the government has a strong determination to solve China’s complex water problems. However, with a relatively constant water supply, increased water demand will have to be met mainly through water savings and improved water quality (Feng et al., 2010; Yu et al., 2010). China has been facing severe water resource management challenges: water scarcity, water pollution and flooding. China’s water scarcity is characterized by insufficient quantities and poor water quality.

1318

Unprecedented rapid economic development combined with population growth and urbanization triggered the conflict between water supply and demand. Historic improper water resources management further intensified these challenges, and potential climate change will make the conflict even worse in future. If these three challenges cannot be addressed properly, China will have to face even more serious challenges in a near future. Effective water resource management, using lessons from around the world, including soft practices and policy changes, can help tackle the challenges. Several recommendations should be seriously considered by the Chinese government: (1) Integrated water resources management based on river basins needs to be strengthened by giving more power to the commissions of major rivers and lakes; (2) A modern water rights system needs to be developed which will hopefully enable water users to trade their water rights and thus improve the efficiency of water usage; (3) Supply-driven water management needs to be changed and future water management needs to pay more attention to water demand management via improving water reuse rates and water conservation; (4) More money needs to be invested to address the water pollution problem in terms of developing more water treatment facilities; (5) To provide more water related information to the public, will increase public participation and develop a “human-water harmony” and water conservation society. Ma and Tao (2010) pointed out that transparency of information, engaging business/industry as partners in environmental collaboration, and integrating non-governmental organisations and civil society organisations with the country’s environmental governance structure would be useful advances to facilitate water conservation. (6) NGOs could play an important role in terms of supervision, increasing public awareness or even solving some water related problems; (7) The future water price needs to reflect the real value of water as a resource, a commodity, and the value of water treatment costs. Addressing China’s water resource management challenges require a holistic, integrated, scientific (including social and economic science) approach with long-term, coordinated efforts. Recently the Chinese government has announced a mandate for integrated water management to transform the country and tackle the grand challenges for water in China. This policy is to be praised for its level of ambition; but it is also absolutely necessary. References An S., Wang Z., Zhou C., Guan B., Deng Z., Zhi Y., Liu Y., Xu C., Fang S., Xu Z., (2006), The headwater loss of the western plateau exacerbates China's long thirst, AMBIO: A Journal of the Human Environment, 35, 271-272. An Y.L., Zhang L.Y., Liu N., Zhou A.X., Zhang T.D., An Y.K., (2012), Field scale analysis on structural changes of microbial community and its relationships


with environmental factors in nitrobenzenecontaminated groundwater during air sparging remediation, Environmental Engineering and Management Journal, 11, 1687-1696. Bailey A., (2010), Population geographies and climate change, Progress in Human Geography, 12, doi: 10.1177/0309132510383358. CAS, (2007), China Sustainable Development Strategy Report 2007: Water Governance and Innovation, Scientific Press, Beijing, China. Cai Y.P., Huang G.H., Tan Q., Chen B., (2011), Identification of optimal strategies for improving ecoresilience to floods in ecologically vulnerable regions of a wetland, Ecological Modelling, 222, 360-369. Chan F.K.S., Mitchell G., Adekola O., McDonald A., (2012), Flood risk in Asia’s urban mega-deltas, Environment and Urbanization Asia, 3, 41-61. Chan S.N., Lee J.H.W., (2010), Impact of river training on the hydraulics of Shenzhen river, Journal of Hydroenvironment Research, 4, 211-223. Chen Q., Kang C., Xia Q., Guan D., (2011), Primary exploration on low-carbon technology roadmap of china's power sector, Energy, 36, 1500-1512. Chen Y.D., Zhang Q., Xu C.Y., Lu X., Zhang S., (2010), Multiscale streamflow variations of the Pearl River basin and possible implications for the water resource management within the Pearl River Delta, China, Quaternary International, 226, 44-53. Cheng X., (2006), Recent progress in flood management in China, Irrigation and Drainage, 55, S75-S82. Chinese Academy of Sciences, (2007), China Sustainable Development Strategy Report 2007 - Water: Governance and Innovation (in Chinese), Science Press, Beijing. Chong W.H.B., Guan D., Guthrie P., (2012), Comparative analysis of carbonization drivers in China’s megacities, Journal of Industrial Ecology, 16, 564-575. Deng J., Jin X., (2000), Study on fishery biodiversity and its conservation in Laizhou Bay and Yellow River estuary, Zoological Research, 21, 76-82. Fan H., Huang H., Zeng T., (2006), Impacts of anthropogenic activity on the recent evolution of the Huanghe (Yellow) River Delta, Journal of Coastal Research, 22, 919-929. Feng K., Hubacek K., Guan D., (2009), Lifestyles, technology and CO2 emissions in China: A regional comparative analysis, Ecological Economics, 69, 145154. Feng K., Hubacek K., Siu Y., Guan D., (2012), Analyzing drivers of regional carbon dioxide emissions for China, Journal of Industrial Ecology, 16, 600-611. Feng K., Hubacek K., Siu Y., Guan, D., (2012), Assessing regional virtual water flows and water footprints in the Yellow River Basin, China: a consumption based approach, Applied Geography, 32, 691-701. Fleming G., (2002), Learning to live with rivers - the ICE's report to government, Proceedings of the Institution of Civil Engineers-Civil Engineering, 150, 15-21. Fraser E., Termansen M., Sun N., Guan D., Dodds P., Feng K., Yu Y., (2008), Quantifying socioeconomic characteristics of drought-sensitive regions: Evidence from Chinese provincial agricultural data, Comptes Rendus Geosciences, 340, 3572-3577. Fu G., Chen S., Liu C., Shepard D., (2004), Hydro-climatic trends of the Yellow River basin for the last 50 years, Climatic Change, 65, 149-178. Fuchs R., Conran M., Louis E., (2011), Climate change and Asia’s coastal urban cities, Environment and Urbanization Asia, 2, 13-28.

Ge Y., Xu W., Gu Z.H., Zhang Y.C., Chen L., (2011), Risk perception and hazard mitigation in the Yangtze River Delta region, China. Natural Hazards, 56, 633-648. Geng Y.I., Zhang J., Zhou Q., Xu C.D., Zhao Y.Q., (2011), Fuzzy synthetic evaluation of weihe water quality, Environmental Engineering and Management Journal, 10, 1477-1484. Gersonius B., Veerbeek W., Subhan A., Stone K., Zevenbergen C., (2011), Toward a More Flood Resilient Urban Environment: The Dutch Multi-level Safety Approach to Flood Risk Management, In: Resilient Cities, Otto-Zimmermann K. (Ed.), Springer Netherlands, Amsterdam. Gu C., Han S., (2010), Climate change and China’s mega urban regions, Frontiers of Architecture and Civil Engineering in China, 4, 418-430. Gu C., Hu L., Zhang X., Wang X., Guo J., (2011), Climate change and urbanization in the Yangtze River Delta, Habitat International, 35, 544-552. Guan, D., Hubacek, K., (2007), Assessment of regional trade and virtual water flows in China, Ecological Economics, 61, 159-170. Guan D., Hubacek K., (2008), A new and integrated hydroeconomic accounting and analytical framework for water resources: A case study for North China, Journal of Environmental Management, 88, 13001313 Guan D., Hubacek K., Weber C.L., Peters G.P., Reiner D., (2008), The drivers of Chinese CO2 emissions from 1980 to 2030, Global Environmental Change: Human and Policy Dimensions, 18, 626-634. Guan D., Peters G.P., Weber C.L., Hubacek K., (2009), Journey to world top emitter – an analysis of the driving forces of China’s recent emissions surge, Geophysical Research Letters, 36, L04709, doi:10.1029/2008GL036540. Harvey H., Hall J.W., Peppe R., (2009), Outline of a Framework for Systematic Decision Analysis in Flood Risk Management, In: 18th World IMACS / MODSIM Congress, Cairns, Australia 13-17 July 2009, University Western Australia, Nedlands, 4290-4297, On line at: http://mssanz.org.au/modsim09. Hu R.L., Yue Z.Q., Wang L.C., Wang S.J., (2004), Review on current status and challenging issues of land subsidence in China, Engineering Geology, 76, 65-77. Huang H.J., Fan H., (2004), Monitoring changes of nearshore zones in the Huanghe (Yellow River) delta since 1976, Oceanologia et Limnologia Sinica, 35, 313-321. Hubacek K., Guan D., Barrett J., Wiedmann T., (2009), Environmental implications of urbanization and lifestyle change in China: Ecological and Water Footprints, Journal of Cleaner Production, 17, 12411248. Hubacek K., Guan D., Barua A., (2007), Changing lifestyles and consumption patterns in developing countries: A scenario analysis for China and India, Futures, 39, 1084-1096. Jiang Y., (2009), China's water scarcity, Journal of Environmental Management, 90, 3185-3196. Khan S., Hanjra M.A., Mu J., (2009), Water management and crop production for food security in China: A review, Agricultural Water Management, 96, 349-360. Kreibich H., Seifert I., Thieken A., Lindquist E., Wagner K., Merz B., (2011), Recent changes in flood preparedness of private households and businesses in Germany, Regional Environmental Change, 11, 59-71. Lee B.Y., Wong W.T., Woo W.C., (2010), Sea-Level Rise and Storm Surge - Impacts of Climate Change on

1319


Hong Kong, Hong Kong Observatory, Hong Kong Special Administrative Region Government, HKIE Civil Division Conference 2010, 12-14 April 2010, Hong Kong, On line at: http://203.129.68.38/publica/reprint/r915.pdf. Lin C., Su J., Xu B., Tang Q., (2001), Long-term variations of temperature and salinity of the Bohai Sea and their influence on its ecosystem, Progress of Oceanography, 49, 7-19. Liu C.M., He X.W., (2000), Water demand predication in the first half of 21st century, China Water Resources, 1, 19-20 Liu J.G., Raven P.H., (2010), China’s environmental challenges and implications for the world, Critical Reviews in Environmental Science and Technology, 40, 823–851. Liu J., Yang G.W., (2012), Water sustainability for China and beyond, Science, 337, 649-650. Lu Z.P., Yao Y.W., (2006), Typhoon disasters in zhejiang province and the establishment of emergency response mechanism, Journal of Catastrophology, 1, 69-71. Luo X.L., Zeng E.Y., Ji R.Y., Wang C.P., (2007), Effects of in-channel sand excavation on the hydrology of the Pearl River Delta, China, Journal of Hydrology, 343, 230-239. Ma J., Tan X., Zhang N., (2010), Flood management and flood warning system in China, Irrigation and Drainage, 59, 17-22. Ma X., Tao J., (2010), Cross-border environmental governance in the Greater Pearl River Delta (GPRD), International Journal of Environmental Studies, 67, 127-136. Marchi L., Borga M., Preciso E., Gaume E., (2010), Characterisation of selected extreme flash floods in Europe and implications for flood risk management, Journal of Hydrology, 394, 118-133. McGranahan G., Balk D., Anderson B., (2007), The rising tide: assessing the risks of climate change and human settlements in low elevation coastal zones, Environment and Urbanization, 19, 17-37. McLean S., Watson P., (2009), A practical approach to development of housing on floodplain land in the UK, Journal of Building Appraisal, 4, 311-320. Meyer V., Scheuer S., Haase D., (2009), A multicriteria approach for flood risk mapping exemplified at the Mulde river, Germany, Natural Hazards, 48, 17-39. Mimura N., Kawaguchi E., (2011), Responses of Coastal Topography to Sea-level Rise, Proceedings of the International Conference on Coastal Engineering, North America, 1, aug. 2011, On line at: http://journals.tdl.org/ICCE/article/view/5309. Minx J.C., Baiocchi G., Peters G.P., Weber C.L., Guan D., Hubacek K., (2011), A carbonizing dragon: China’s fast growing CO2 emissions revisited, Environmental Science & Technology, 45, 9144–9153. MWR, (2007a), Water Resources Bulletin 2006, Ministry of Water Resources, P.R. China, Beijing, China. MWR, (2012), Water Resources Bulletin 2011, Ministry of Water Resources, P.R. China, Beijing, China. MWR, (2007b), The 11th Five-Year Plan of National Water Resources Development, Ministry of Water Resources, P.R. China, Beijing, China. Ng C.N., Xie Y.J., Yu X.J., (2011), Measuring the spatiotemporal variation of habitat isolation due to rapid urbanization: A case study of the Shenzhen River cross-boundary catchment, China, Landscape and Urban Planning, 103, 44-54.

1320

Nicholls R., Hanson C., Herweijer N., Patmore S., Hallegatte C.J., Morlot, J.C., Wood R.M., (2008), Ranking Port Cities with High Exposure and Vulnerability to Climate Extremes: Exposure Estimates, In: OECD Environment Working Paper, No. 1, OECD Publishing, DOI:10.1787/19970900. Nicholls R.J., (1995), Coastal megacities and climate change, GeoJournal, 37, 369-379. Pitt M., (2008), Learning Lessons from the 2007 Floods, Final Report, The Cabinet Office, London, UK, On line at: http://webarchive.nationalarchives.gov.uk/+/http:/ww w.cabinetoffice.gov.uk/upload/assets/www.cabinetoffi ce.gov.uk/flooding_review/flood_report_web.pdf. Plate E.J., (2007), Early warning and flood forecasting for large rivers with the lower Mekong as example, Journal of Hydro-environment Research, 1, 80-94. Qu J.H., Fan M.H., (2010), The current state of water quality and technology development for water pollution control in China, Critical Reviews in Environmental Science and Technology, 40, 519–560. Research team of china climate change country study, (1999), Climate Change Country Study, China (in Chinese), Beijing, Tsinghua University Press. Seto K.C., (2011), Exploring the dynamics of migration to mega-delta cities in Asia and Africa: Contemporary drivers and future scenarios, Global Environmental Change, 21, S94-S107. Shi P.J., Yuan Y., Zheng J., Wang J.A., Ge Y., Qiu G.Y., (2007), The effect of land use/cover change on surface runoff in Shenzhen region, China, Catena, 69, 31-35. SEPA, (2007), China Environmental Bulletins 1990–2006 (in Chinese), State Environmental Protection Administration, Beijing, China. SOA, (2013), China ocean disasters Bulletin 2012, State Oceanic Administration, P. R. China, Beijing, China State Council, (2006), The 11th Five-Year Plan of the Prople’s Republic of China, Government of the People’s Republic of China, Beijing, China. State Council, (2011), The 12th Five-Year Plan for National Environmental Protection of the People's Republic of China, Government of the People’s Republic of China, Beijing, China, On line at: http://english.mep.gov.cn/Plans_Reports/12plan/20120 1/P020120110355818985016.pdf. Syvitski J., (2008), Deltas at risk, Sustainability Science, 3, 23-32. Syvitski J., Kettner A.J., Overeem I., Hutton E.W.H., Hannon M.T., Brakenridge G.R., Day J., Vorosmarty C., Saito Y., Giosan L., Nicholls R.J., (2009), Sinking deltas due to human activities, Nature Geoscience, 2, 681-686. Tang Y.Q., Cui Z.D., Wang J.X., Yan L.P., Yan X.X., (2008), Application of grey theory-based model to prediction of land subsidence due to engineering environment in Shanghai. Environmental Geology, 55, 583-593. UN-HABITAT., (2008), State of the World's City Report, Nairobi, United Nations. Varis O., Biswas A.K., Tortajada C., Lundqvist J., (2006), Megacities and water management, International Journal of Water Resources Development, 22, 377394. Varis O., Vakkilainen P., (2001), China's 8 challenges to water resources management in the first quarter of the 21st Century, Geomorphology, 41, 93-104. Wang X.C., Guo H.Y., (2000), Hai Water Resource Facing the 21st Century, Tianjin, China, Tianjin Science and Technology Press.


Wang X., Jin P., (2006), Water Shortage and Needs for Wastewater Re-Use in The North China, IWA Publishing, London, United Kingdom. Wang L., Zheng Z., Luo X., Zhang J., (2010), The current pollution status and control technology of the Taihu Lake Basin, Jiangsu Province, China, International Journal of Environmental Studies, 67, 195-205. White I., Kingston R., Barker A., (2010), Participatory geographic information systems and public engagement within flood risk management, Journal of Flood Risk Management, 3, 337-346. World Bank, (2007), Addressing Water Scarcity in China. Recommendations for Selected Water Resource Management Issues, World Bank, Washington, On line at: http://wwwwds.worldbank.org/external/default/WDSContentServ er/WDSP/IB/2009/01/14/000333037_2009011401112 6/Rendered/PDF/471110PUB0CHA0101OFFICIAL0 USE0ONLY1.pdf Xia J., Chen Y. D., (2001), Water problems and opportunities in the hydrological sciences in China, Hydrological Sciences Journal, 46, 907-921. Xia J., Liu M., Jia S., Song X., Luo Y., Zhang S., (2004), Water security problem and research perspective in North China, Journal of Natural Resources (in Chinese), 19, 550-560. Xia J., Zhang L., Liu C., Yu J., (2007), Towards better water security in North China, Water Resources Management, 21, 233-247. Xu Y.S., Zhang D.X., Shen S.L., Chen L.Z., (2009), Geohazards with characteristics and prevention measures

along the coastal regions of China, Natural Hazards, 49, 479-500. Yang D., Li C., Hu H., Lei Z., Yang S., Kusuda T., Koike T., Musiake K., (2004), Analysis of water resources variability in the Yellow River of China during the last half century using historical data, Water Resources Research, 40, 1-12. Yeung, Y.M., (2010), The further integration of the Pearl River Delta, Environment and Urbanization Asia, 1, 13-26. Yeung, Y.M., (2011), Rethinking Asian cities and urbanization: four transformations in four decades, Asian Geographer, 28, 65-83. Yu X., (2010), China’s environment and natural resource management, International Journal of Environmental Studies, 67, 115-116. Zhang G.L., (1999), Water Supply and Demand in China in the 21st Century (in Chinese), Chinese Water Resources and Hydropower Press, Beijing, China. Zhang J., (2009), A vulnerability assessment of storm surge in Guangdong Province, China, Human and Ecological Risk Assessment: An International Journal, 15, 671-688. Zhang Q., Zhang W., Chen Y., Jiang T., (2011), Flood, drought and typhoon disasters during the last halfcentury in the Guangdong province, China, Natural Hazards, 57, 267-278. Zhou X., Cai L., (2010), Coastal and marine environmental issues in the Pearl River Delta region, China., International Journal of Environmental Studies, 67, 137-145.

1321

PROJECT PRESENTATION

ADVANCED SEPARATION OF BIOSYNTHETIC COMPOUNDS BY FACILITATED AND SYNERGETIC PERTRACTION (PERSYNBIO) Research Grant no. 207/5.10.2011, Project ID: PN-II-IDEI-PCE-2011-3-088 (http://www.ch.tuiasi.ro/cercetare/IDEI/dancascaval/asbcfsp/Idei-proiect2.htm) The proposed research project, namely the study on separation by free, facilitated and/or synergetic pertraction (extraction through liquid membranes) of some valuable biosynthetic products, joins the top of the worldwide researches on separation of biosynthetic compounds with important applications in medicine, food and cosmetics, researches that are connected to the current context of “the white biotechnology”. Compared to the chemical methods, the biosynthesis represents a very advantageous alternative for production of many compounds with biological activity, because of the reduction of the overall process stages number and of the advanced utilization of the lowcost raw materials. However, the undesirable particularity of industrial biotechnologies is the complexity of the separation from fermentation broths of the obtained products, especially due to their high dilution in broth, chemical and thermal lability and to the presence of secondary products. Therefore, the purification of biosynthetic compounds requires a laborious succession of separation stages with high material and energy consumption, the contribution of these stages to the overall cost being of 20 - 60%, or even more. For this reason, most of the research and development directions on separation techniques are focused on the biosynthetic compounds. In this context, among the wideworld studies on new separation methods for the biosynthetic products two major directions can be pointed out: reactive extraction and extraction through liquid membranes, or pertraction. Extraction and transport through liquid membranes, also called pertraction or permeation, represents the development of reactive extraction technique. Pertraction consists in the transfer of a solute between two aqueous phases of different pH or other chemical properties value, phases that are separated by a solvent layer of various sizes. By comparing the extraction using liquid membranes with conventional liquid-liquid extraction, the advantages of the pertraction are as follows: the quantity of solvent used is small, because of its continuous regeneration; the loss of solvent during extraction and transport process is reduced; as long as the pH gradient between the two aqueous phases is maintained, there is the possibility of solute transport against its concentration gradient; higher diffusion coefficient of permeates in liquid membranes compared to polymeric membranes; energy consumption is very low. The pertraction efficiency could be significantly enhanced by adding one or more carriers in liquid membrane, such as organophosphoric compounds, long chain amines or crown-ethers, the separation process being called facilitated pertraction (for one carrier) or synergetic pertraction (for mixture of carriers or solvents inducing the synergetic effect). Because the use of biotechnology and the optimization of the high costs, biotechnological stages represent one of the priorities, the proposed project contributes to the development of the fundamental and

applicative researches in the bioengineering and biotechnology domain in our country, all the more as the activity and the previous results of our team in this field are appreciated in the scientific media. New valences of the multidisciplinary research are promoted by means of this project, in the purpose of increasing the economically efficiency of the biotechnologies for production of antibiotics, carboxylic acids, amino acids and vitamins with pharmaceutical, food and cosmetic utilizations. Therefore, the original objectives of the proposed project and the novelty and scientifically interest of the corresponding potential results are as follows (the potential results could be considered as intermediate milestones): 1. individual and selective separation of carboxylic acids obtained by succinic fermentation with Actinobacillus succinogenes (succinic, formic and acetic acids) by facilitated pertraction; 2. individual and selective separation of carboxylic acids obtained by propionic fermentation with Torulopsis glabrata (propionic, pyruvic acids, and acetic acid – the last acid included in the above 3. objective) by facilitated pertraction; 4. separation by reactive extraction and facilitated pertraction of mupirocin (pseudomonic acid, antibiotic produced by Pseudomona fluorescens); 5. separation of nistatine by free or facilitated pertraction from unfiltered fermentation broths of Streptomyces; 6. separation of cinnamic and p-methoxycinnamic acids by synergetic reactive extraction and synergetic facilitated pertraction; 7. separation by reactive extraction and facilitated pertraction of vitamin B5 (pantothenic acid); 8. separation by synergetic facilitated pertraction of vitamin B9 (folic acid). The studied biosynthetic compounds are products of high economically value, as it was indicated in the studies for 2011 and previsions for the next years made by specialized companies (Market News Service, In-Pharma Technologist, Nutra Ingredients Europe, Codex Alimentarius Commission). Besides the above mentioned objectives, another major objective of the project is to reach top-results in the field of bioseparations, for increasing the visibility of the Romanian research activities, and for translating the obtained results to the industrial practice. The project team includes two senior researchers, three postdoctoral researchers and three doctoral researchers.

Project leader: Professor Dan Caşcaval, Ph.D Department of Organic, Biochemical and Food Engineering “Gheorghe Asachi” Technical University of Iasi [email protected], [email protected]


July 2013, Vol.12, No. 7, 1323-1328



RADIATION-INDUCED DECOMPOSITION AND POLYMERIZATION OF POLYVINYL ALCOHOL IN AQUEOUS SOLUTIONS Weihua Sun1, Lujun Chen1,2, Jinping Tian1, Jianlong Wang3 , Shijun He3 1 School of Environment, Tsinghua University, Beijing, P.R. China Zhejiang Provincial Key Laboratory of Water Science and Technology, Zhejiang Province, Jiaxing, P.R. China 3 Institute of Nuclear and New Energy Technology, Tsinghua University, Beijing, P.R. China

2

Abstract Polyvinyl alcohol (PVA) is a typical refractory organic pollutant with low biodegradability and high molecular weight. This paper presents data on the radiolysis of PVA in aqueous solutions by using ionizing radiation. Response surface methodology with Box-Benhken design of experiment was applied to analyze the interactions among initial concentration of PVA, absorbed dose, and initial pH of aqueous solutions on PVA removal efficiency. 9.4-89.4% PVA was removed based on different combination conditions. In addition, absorbed dose and its interactions with PVA initial concentration were determined to be significant in PVA degradation, and both acidic and alkaline conditions were more beneficial to PVA degradation than neutral condition. When PVA initial concentration was 200 mg/L and absorbed dose was 2.75 kGy, PVA removal ratio was obtained to yields of 88.8% and 89.4% at pH 1 and pH 13. Interestingly, polymerization and gelation of PVA in aqueous solution was found during the irradiation, and gel ratio was 67.2% at 9 kGy when PVA initial concentration was 3000 mg/L. Finally, individual effect of hydroxyl radicals, hydrated electrons and hydrogen atoms on PVA radiolysis was studied, respectively. Key words: gamma radiation, ionizing radiation, microfiltration, polymerization, PVA Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Polyvinyl alcohol (PVA), a water-soluble polymer with properties of high strength and excellent film forming, is widely used as a textile size, grease-resistant paper coatings and adhesives (Patachia et al., 2011). However, typical PVA wastewater discharged from textile wastewater is generally refractory biodegradable (Chen et al., 2000). The global annual consumption of PVA had been over one million tons (Inoguchi and Chinn, 2010). The large amount of discharged PVA from industrial effluents has become a significant pollution problem. Once PVA wastewater discharge into nature waters without being treated efficiently, the activities of aerobic microorganisms will be inhibited caused by water surfaces foaming (Shan et al., 2009) and water viscosity increasing which is detrimental to 

ecosystem. In addition, PVA could be lead to accumulation in human body through food chains (Oh et al., 2009). Even worse, conventional wastewater plants are difficult to degrade PVA due to its poor biodegradable. Therefore, effective methods for PVA removal in wastewaters need to be researched. PVA wastewater treatment has been focused on filtration technologies, biological processes and advanced oxidation processes. Filtration technologies, i.e., microfiltration (Zhang et al., 2007), ultrafiltration ( Lee et al., 1999; Lin et al., 1995; Lin et al., 1997; Porter, 1998) and nanofiltration (Mo et al., 2008) were usually used for PVA recovery. However, PVA film forming on membrane surface and membrane blocking baffled its application. Biological process with a long treating period was always affected by water quality fluctuation (Ciner et

Author to whom all correspondence should be addressed: E-mail: [email protected], Phone: +86 010 62776686, Fax: +86 010 62797581

Sun et al./Environmental Engineering and Management Journal 12 (2013), 7, 1323-1328

al., 2003; Choi et al., 2004; Liu et al., 2010; Patachia et al., 2011; Qian et al., 2004). Compared with other technologies, advanced oxidation processes (AOPs) with properties of short treatment time, good performance on removal efficiency and widely environmental application, have been proved to be an effective way for PVA removal (Betianu et al., 2008; Chen et al., 2000; Lin et al., 1997; Sun et al., 2004; Wu et al., 2008; Xia et al., 2000). Ionizing radiation is a clean and sustainable advanced oxidation technology without adding chemicals and producing dangerous wastes. As Eq. (1) shows, primary active radicals, i.e., hydroxyl radicals (•OH), hydrated electrons (eaq−) and hydrogen atoms (H•) are generated through water radiolysis induced by gamma ray or electron beam (Wasiewicz et al., 2006). These high-energy particles could decompose refractory macromolecular contaminants to small molecules with low-toxic, nontoxic, or even be mineralized completely.  H 2O  2.8HO   2.7eaq  0.6H  

 0.7H 2O2  0.45H 2  3.2H aq  0.5HOaq (1) Previously, little research was reported on treatment of PVA-containing wastewater by using ionizing radiation. Zhang and Yu (2004) documented that PVA radiolysis was initiated by •OH and H•, and peroxyl radicals played minor role for PVA degradation when oxygen existed. As Hu and Wang (2009) documented gamma radiation combined with hydrogen peroxide was a useful method for improving •OH yield. In addition, Zhang et al. (2005) proved that gamma radiation combined with hydrogen peroxide was a useful method for improving PVA removal efficiency. However, excess H2O2 had been noted to play a role as •OH scavenger, making the PVA degradation process less effective (Zhang and Yu, 2004; Matilainen and Sillanpaa, 2010). Interestingly, PVA gel was found during gamma radiation (Danno, 1958). This phenomenon required sufficient radiation dose and higher PVA initial concentration. Chen et al. (1985) found that the crosslink of PVA aqueous solutions was initiated by •OH and H•, but not the species eaq−. To found the relationship of decomposition and polymerization of PVA during ionizing radiation, this study focused on the combination effects of the main factors on PVA removal in aqueous solutions by using gamma radiation at high dose rate which was more close to practical application.

In addition, response surface methodology, an excellent tool for optimizing experimental conditions (Moghaddam et al., 2010; Prabhakaran et al., 2010), was used for experimental design, data analysis, and model building. 2. Materials and methods PVA (the average polymerization degree = 1700  200, chemical pure), Iodine (analytical pure, AP), Potassium iodide (AP), Boric acid (AP) were purchased from commercial channels. Reagents and chemicals were used without further purification. All samples were prepared in deionized water. Gamma radiation was carried out in a 60Co source (Tsinghua University, Beijing, China). The dose rate used in this study was 158 Gy/min. PVA solutions were prepared and irradiated in radiation-proof glass tubes ( = 20 mm). In addition, PVA analyses were carried out by a UVVis spectrophotometer (UV-3100, Shimadzu Co.) according to the procedure described by Finley (1961). PVA gel fraction ratio was measured by constant weight method described by Sun (2012). pH measurements were performed with a pH meter (METTER TOLEDE, Switzerland), and pH adjustment by sodium hydroxide solution and sulfuric acid solution. All operations were performed at ambient temperature.Initial concentration of pollutant, absorbed dose and initial pH of aqueous solution are key factors in ionizing radiation. Pollutant’s structure and concentration determine levels of absorbed dose which has positive correlation with process cost and pollutant removal efficiency in most cases of radiation decomposition. Furthermore, pH affects production fields of radicals in water radiolysis. Based on actual printing and dyeing wastewater, PVA initial concentration range was chose to 50-200 mg/L in this study. Compare with the previous reports (Zhang et al., 2004; Zhang et al., 2005), the values of absorbed dose were chose to 0.5-5.0 kGy. In order to judge radicals affection clearly, strong acidic neutral and strong alkaline conditions were determined finally. For experimental design, data analysis, and model building, the Design Expert 8.0.0 software (Stat-Ease Inc., Minneapolis, Minn., U.S.A.) was employed. A Box-Behnken design (BBD) with 3 variables was used to determine the response pattern and then to establish a model. 3 variables used in this study were PVA initial concentration (X1), absorbed dose (X2), and initial pH value (X3), with 3 levels of each variable (Table 1).

Table 1. Experimental ranges and levels of the independent variables

1324

Factor

Unit

Symbol

Initial concentration Absorbed dose Initial pH

mg/L kGy -

X1 X2 X3

-1 50 0.5 1.0

Level 0 125 2.75 7.0

1 200 5.0 13.0

Radiation-induced decomposition and polymerization of polyvinyl alcohol in aqueous solutions

Values of "Prob >F" less than 0.05 indicated model terms were significant. In this study, X2, X1X2, X12, X22 were significant model terms. Therefore, the most important parameters which affected PVA removal by gamma radiation were the primary effect of absorbed dose, the interaction between absorbed dose with PVA initial concentration, and the quadratic effects of concentration and absorbed dose, respectively. Using “Numerical Optimization” of the software, values of initial concentration, absorbed dose and initial pH were set for their full ranges, respectively. Then the goal for PVA removal efficiency (90-100%) was set under the “Solutions” branch. Through analysis of the first 30 solutions found by numerical optimization, the conditions for the optimal ranges (PVA removal efficiency > 90%) were obtained: (1) under the acidic conditions (pH = 1.1-2.5), the optimal range of PVA initial concentration was 133-175 mg/L, and the optimal range of absorbed dose was 3.4-4.5 kGy; (2) under the alkaline conditions (pH = 10.3-13.0), the optimal range of PVA initial concentration was 130-176 mg/L, and the optimal range of absorbed dose was 3.3-4.5 kGy. However, the optimal range didn’t been found under the neutral conditions, indicated that acidic or alkaline condition was more beneficial to PVA degradation than neutral condition.

3. Results and discussion 3.1. Model analysis Based on the results of BBD design and PVA removal efficiency (the dependent variable Y, Table 2), the response surface quadratic model was established. Model terms were evaluated based on the P value (the probability of obtaining a test statistic) with 95% confidence level. The results were completely analysed using analysis of variance (ANOVA) by Design Expert software. The constant and coefficients of response equation was obtained from the software analysis (Table 3). The equation of actual factors shows as following (Eq. 2): Y = 11.7767 + 0.51941X1 +19.0617X2 - 2.50324X3 + 0.08385X1X2 - 0.00044X1X3 - 0.02963X2X3 0.00262X12-3.80247X22+0.19722X32 (2) Three-dimensional plots and their respective contour plots were obtained based on the effect of the levels of the three factors. The ANOVA analysis for response surface quadratic model was presented in Table 3. Model F-value of 5.64 implied the model was significant. There was only a 1.6% chance that a "Model F-Value" this large could occur due to noise. In addition, a negative predicted R2=0.938 implied that the overall mean was a better predictor. Adeq Precision measures the signal to noise ratio. A ratio greater than 4 was desirable. The ratio of 7.603 indicated an adequate signal. This model could be used to navigate the design space.

3.2. Influence factors on PVA radiolysis Fig. 1 shows the effect of different single factor on PVA maximum removal efficiency.

Table 2. The Box-Behnken design and the values of response Run Number X1 0 0 -1 0 0 0 1 0 1

1 2 3 4 5 6 7 8 9

Levels X2 0 0 0 1 0 0 1 -1 -1

Y/% X3 0 0 -1 -1 0 0 0 1 0

79.4 79.4 53.7 83.1 79.4 79.4 77.4 52.2 6.10

Run Number 10 11 12 13 14 15 16 17

X1 1 1 -1 0 0 -1 0 -1

Levels X2 0 0 0 0 -1 1 1 -1

Y/% X3 -1 1 1 0 -1 0 1 0

88.8 89.4 55.1 79.4 49.5 56.4 84.2 41.7

Table 3. ANOVA analysis for response surface quadratic model and final equation in terms of factors Source Model Intercept X1 X2 X3 X1X2 X1X3 X2X3 X12 X22 X32

Sum of Squares

df

6778.50 375.38 2872.82 4.21 800.89 0.16 0.64 916.05 1560.26 212.25

9 1 1 1 1 1 1 1 1 1

Mean Square

F Value

753.17 375.38 2872.82 4.21 800.89 0.16 0.64 916.05 1560.26 212.25

5.64 2.81 21.52 0.032 6.00 0.0012 0.0048 6.86 11.69 1.59

p-value (Prob > F) 0.0163 0.1375 0.0024 0.8642 0.0441 0.9733 0.9467 0.0344 0.0112 0.2477

Constant and Coefficients β0 β1 β2 β3 β12 β13 β23 β11 β22 β33

Estimate value Coded Actual Factors Factors 79.40 +11.7767 6.85 +0.51941 18.95 +19.0617 0.73 -2.50324 14.15 +0.08385 -0.20 -0.00044 -0.40 -0.02963 -14.75 -0.00262 -19.25 -3.80247 7.10 +0.19722

Standard Error

VIF

5.17 4.08 4.08 4.08 5.78 5.78 5.78 5.63 5.63 5.63

1.00 1.00 1.00 1.00 1.00 1.00 1.01 1.01 1.01

1325


X1

100

a

PVA Maximum Removal/%

80 60

X2

100 80

b

60 100

X3

c

80 60 -1 .0

-0 .6

-0 .2 0 .2 L e v e ls

0 .6

1 .0

Fig. 1. Effect of single factors on the PVA maximum removal efficiency

Using “Numerical Optimization” of the software, the variation of maximum removal efficiency following the initial concentration at different levels with respect to absorbed dose and initial pH in their full ranges was obtained. The maximum fluctuation value-a caused by initial concentration was found to be 26% (Fig. 1a). Through the same analysis method, the maximum fluctuation values caused by absorbed radiation and by initial pH were found to be b = 41% and c = 3%, respectively (Fig. 1b-c). Thus, due to the value order of b > a > c, the factors affected on PVA maximum removal efficiency followed the order: absorbed dose > initial concentration > initial pH. The response surfaces of PVA removal are shown in Fig. 2-3. Each of three-dimensional models for PVA removal respected to two factors within the design space with the third factor at the middle level. As shown in Fig. 2, when PVA initial concentration was at a low level, e.g., 50 mg/L, the PVA removal efficiency is firstly increased, then decreased with the absorbed dose increasing. Under gamma radiation, soluble monomers and polymers in aqueous solution could be converted into higher molecular weight substances (Getoff, 1996). In addition, hydrogenabstracted PVAs could recombine (Zhang et al., 2005). Therefore, PVA polymerization including PVA recombination existed except PVA decomposition in this study. Thereafter, the radiolysis process of PVA could be described by Eq. (3). At any time t the removed PVA (PVARemoval) was equal to the sum of decomposed PVA (PVAD) and polymerized PVA (PVAP) minus recombined PVA (PVAR). The total removed PVA was equal to the integral from zero time to t time of PVARemoval. If PVARemoval < 0 during a period time, e.g., Δt = t1 – t, the removal rate appeared to decrease. When PVA initial concentration was at a relative high level, e.g., 200 mg/L, the PVA removal efficiency always increased with the absorbed dose increasing which indicated PVARemoval was always above zero.

1326

PVARemoval = PVAD + PVAP – PVAR

(3)

The relationship of absorbed dose and initial pH on PVA degradation is presented in Fig. 3. Whenever how absorbed dose levels changed, PVA removal efficiency always decreased at first, and then increased with the pH value increasing. Thus, PVA removal was more effective under acidic or alkaline conditions than that under neutral conditions. At acidic conditions, γ-lactone was the form of a terminal group of PVA. The combination of proton and lone pair electron formed a new bond destroyed the ester bond in the acidolysis process. At alkaline conditions, the terminal group of PVA was carboxyl salt. PVA degradation mechanism might be dominated by dynamic process (Xia et al., 2000). Therefore, acid and alkali played a catalytic role in PVA degradation. In addition, pH values affected water radiolysis production yields of the primary radicals. Under strong acidic conditions, eaq− could be converted to H• under radiation. Oppositely, H• was converted to eaq−, and •OH was decomposed to Haq+ and Oaq− under strong alkaline conditions (Getoff, 1996). 3.3. Effect of primary radicals on PVA radiolysis To clarify which radicals played the key role in PVA radiolysis, the isolated experiments of •OH, H• and eaq− were carried out. For determination of the reaction of PVA and •OH, aqueous solutions were sparged with N2O, which converted eaq− and H• to • OH (Yang et al., 2006). To determine the reaction of eaq−, aqueous solutions were pre-saturated with N2 to remove dissolved O2 above pH 3, and adding 0.2 mol/L t-BuOH to scavenge formed •OH and H• (Yang et al., 2006). Reactions with H• atoms were studied in N2 saturated solutions containing 0.2 mol/L t-BuOH below pH 2 (Getoff, 1996). As shown in Table 4, when initial PVA was 300 mg/L, both •OH and H• played the main role in PVA degradation. 98% and 97% PVA were removed by individual •OH and H• at 9 kGy, respectively. However, the reaction with eaq− played little role in PVA degradation until absorbed dose reached to 6 kGy. When absorbed dose was enhanced to 9 kGy, 54% PVA was removed by eaq−. When initial PVA was enhanced to 3000 mg/L, H• also played the main role in PVA degradation, and eaq− also played little role in PVA degradation. But •OH radicals didn’t play any role on PVA degradation before 9 kGy. As part “3.2” concluded, PVA polymerization existed during radiation process. In addition, most radiation polymerizations were radical polymerizations. Furthermore, crosslinking was distinguished by the occurrence of gelation at some point in the polymerization (Odian, 2004). Due to the crosslinking, PVA molecules in the solution could interlink with each other to form an insoluble gel under gamma radiation.

Radiation-induced decomposition and polymerization of polyvinyl alcohol in aqueous solutions

100

80

PVA Removal/%

PVA Removal/%

100 60 40 20 0

1.0

0.6

0.2

X2

-0.2

-0.6

-1.0

-1.0

-0.6

-0.2

0.2

0.6

60 40 20 0

1.0

1.0

0.6

X3

X1

Fig. 2. The 3D model for PVA removal with respect to X1Initial Concentration and X2-Absorbed Dose within the design space with X3-Initial pH=7.0

80

0.2

-0.2

-0.6

-1.0

-1.0

-0.6

-0.2

0.2

0.6

1.0

X2

Fig. 3. The 3D model for PVA removal with respect to X2Absorbed Dose and X3-Initial pH within the design space with X1-Initial Concentration = 125 mg/L

Table 4. Variation of absorbed dose with PVA/PVA0 at different primary radicals conditions PVA0a 300 mg/L PVA0 3000 mg/L

•

OH H• eaq− • OH H• eaq−

0 kGy 1.00 1.00 1.00 1.00 1.00 1.00

1 kGy 0.98 1.05 0.95 1.13 0.90 0.94

3 kGy 0.68 0.70 1.02 1.01 0.81 0.94

6 kGy 0.11 0.13 0.88 1.00 0.81 0.93

9 kGy 0.02 0.03 0.46 0.92 0.71 0.95

a: PVA0 was standard for initial PVA concentration

As shown in Fig. 4, the insoluble gels were observed under a full •OH condition. The gel ratio of PVA was obtained to 67.2% at 9 kGy.

Fig. 4. The effect of different primary radicals on PVA gelation (initial PVA = 3000 mg/L)

dominated the PVA radiolytic process at low PVA concentration, i.e., 0-300 mg/L. The significant primary factor for PVA decomposition was absorbed dose. The interactive effect of absorbed dose and PVA initial concentration also played an important role in PVA decomposition. When PVA initial concentration was 200 mg/L, the maximum removal ratio of PVA was 89.4%. In addition, both hydroxyl radicals and hydrogen atoms were the key radicals in PVA decomposition. Furthermore, polymerization existed in the gamma radiation of PVA in aqueous solutions, and the polymerization could convert to gelation at high PVA initial concentration level, i.e., 3000 mg/L. Also, hydroxyl radicals were found to be the key radicals in PVA polymerization. The PVA gel ratio was 67.2% at 9 kGy under a full hydroxyl radical condition, indicated that ionizing radiation, i.e., gamma radiation combined with physical method, i.e., microfiltration was a feasible method to treat PVA wastewater.

Therefore, •OH radicals played two different roles during PVA radiolysis. At the low PVA initial concentration, the main role of •OH was PVA decomposition. Oppositely, the main role of •OH was PVA polymerization at high PVA initial concentration. As Table 4 and Fig. 4 presented, •OH almost did not played any role in PVA decomposition due to most of •OH radicals were consumed to generate PVA gel. Importantly, the insoluble gel could be removed by microfiltration which indicated ionizing radiation combined physical methods such as microfiltration was a feasible strategy for PVA removal.

This study was accomplished under the National High Technology Research and Development Program (No. 2009AA063905) supported by the Chinese Ministry of Science and Technology and the Independent Research Project (No. 20101081929) supported by Tsinghua University.

4. Conclusions

References

Ionizing radiation was an effective method of PVA removal through by PVA decomposition and PVA polymerization. The decomposition effect

Betianu C., Caliman F.A., Gavrilescu M., Cretescu I., Cojocaru C., Poulios I., (2008), Response surface methodology applied for Orange II photocatalytic degradation in TiO2 aqueous suspensions, Journal of

Acknowledgements

1327


Chemical Technology and Biotechnology, 83, 14541465. Chen G.H., Lei L.C., Yue P.L., Cen P., (2000), Treatment of desizing wastewater containing poly(vinyl alcohol) by wet air oxidation, Industry & Enginnering Chemistry Research, 39, 1193-1197. Chen W.X., Bao H.Y., Zhang M.W., (1985), Effect of gamma radiation on gelation in polyvinyl alcohol solutions, Radiation Physics and Chemistry, 26, 4347. Choi K.K., Park C.W., Kim S.Y., Lyoo W.S., Lee S.H., Lee J.W., (2004), Polyvinyl alcohol degradation by Microbacterium barkeri KCCM 10507 and Paeniblacillus amylolyticus KCCM 10508 in dyeing wastewater, Journal of Microbiology and Biotechnology, 14, 1009-1013. Ciner F., Solmaz S.K.A., Yonar T., Ustun G.E., (2003), Treatability studies on wastewater from textile dyeing factories in Bursa, Turkey, International Journal of Environment and Pollution, 19, 403-407. Danno A., (1958), Gel formation of aqueous solution of polyvinyl alcohol irradiated by gamma rays from Cobalt-60, Journal of the Physical Society of Japan, 13, 722-727. Finley J.H., (1961), Spectrophotometric determination of polyvinyl alcohol in paper coatings, Analytical Chemistry, 33, 1925-1927. Getoff N., (1996), Radiation-induced degradation of water pollutants-state of the art, Radiation Physics and Chemistry, 47, 581-593. Hu J., Wang J., (2009), Effect of H2O2 on the kinetics of radiolysis degradation 4-chlorophenol in aqueous solution, Journal of Tsinghua University (Science and Technology), 49,111-113. Lee S.W., Haam S.J., Kwak J.W., Jang J.H., Lee Y.C., (1999), Ultrafiltration of desizing wastewater containing PVA in bench scale test, Environmental Technology, 20, 277-283. Lin S.H., Lan W.J., (1995), Polyvinyl alcohol recovery by ultrafiltration: effects of membrane type and operating conditions, Separations Technology, 5, 97-103. Lin S.H., Lan W.J., (1997), Semi-batch recovery of polyvinyl alcohol from aqueous solution by ultrafitration, Journal of Dispersion Science and Technology, 18, 463-475. Lin S.H., Lo C.C., (1997), Fenton process for treatment of desizing wastewater, Water Research, 31, 2051-2056. Liu R.R., Tian Q., Yang B., Chen J.H., (2010), Hybrid anaerobic baffled reactor for treatment of desizing wastewater, International Journal of Environment Science and Technology, 7, 111-118. Matilainen A., Sillanpaa M., (2010), Removal of natural organic matter from drinking water by advanced oxidation processes, Chemosphere, 80, 351-365. Mo J.H., Lee Y.H., Kim J., Jeong J.Y., Jegal J., (2008), Treatment of dye aqueous solutions using nanofiltration polyamide composite membranes for the dye wastewater reuse, Dyes and Pigments, 76, 429434. Moghaddam S.S., Moghaddam M.R.A., Arami M., (2010), Coagulation/flocculation process for dye removal using sludge from water treatment plant: Optimization through response surface methodology, Journal of Hazardous Materials, 175, 651-657. Odian G., (2004), Principles of Polymerization, John Wiley & Sons, Hoboken.

1328

Oh S.Y., Kim H.W., Park J.M., Park H.S., Yoon C., (2009), Oxidation of polyvinyl alcohol by persulfate activated with heat, Fe2+, and zero-valent iron, Journal of Hazardous Materials, 168, 346-351. Patachia S., Mincea D., Scarneciu C., (2011), Efficiency of the poly (vinyl alcohol) (PVA) hydrogels as soil conditioner, determined by monitoring the Capsicum sp. L. growth, Environmental Engineering and Management Journal, 10, 225-230. Porter J.J., (1998), Recovery of polyvinyl alcohol and hot water from the textile wastewater using thermally stable membranes, Journal of Membrane Science, 151, 45-53. Prabhakaran D., Basha C.A., Kannadasan T., Aravinthan P., (2010), Removal of hydroquinone from water by electrocoagulation using flow cell and optimization by response surface methodology, Journal of Environmental Science and Health Part aToxic/Hazardous Substances & Environmental Engineering, 45, 400-412. Qian D., Du G.C., Chen J., (2004), Isolation and culture characterization of a new polyvinyl alcohol-degrading strain: Penicillium sp WSH02-21, World Journal of Microbiology & Biotechnology, 20, 587-591. Shan J.C., Guan Y., Zheng Q.K., Han J.S., Liu Q.S., Pu Z.Y., (2009), Application of Urea/H2O2 activationoxidation system in degradation of PVA and desizing of polyester/cotton fabric, Journal of Applied Polymer Science, 113, 860-867. Sun W., Tian J., Chen L., He S., Wang J., (2012), Improvement of biodegradability of PVA-containing wastewater by ionizing radiation pretreatment, Environmental Science And Pollution Research, 19, 3178-3184. Sun Z.S., Yang Y., Chen Y.X., (2004), Photocatalytic degradation of polyvinyl alcohol in UV TiO2-H2O2 system, Acta Energlae Solaris Sinica, 25, 760-763. Wasiewicz M., Chmielewski A.G., Getoff N., (2006), Radiation-induced degradation of aqueous 2,3dihydroxynaphthalene, Radiation Physics and Chemistry , 75, 201-209. Wu Y., Lian Y., Liu J., Zhu H., (2008), Sonochemical degradation of ployvinyl alchohol in aqueous solutions (in Chinese), Technology of WaterTreatment, 34, 3234, 59. Xia L., Li K., Pang J., Cao G., Xi Z., (2000), Polyvinyl alcohol degradation under microwave irradiation (in Chinese), Environment Chemisty, 19, 556-560. Yang C., Wang D., Guo J., Zheng P., (2006), Comparing methods for determination of hydroxyl radicals during advanced oxidation process in water (in Chinese), Techniques and Equipment for Environmental Pollution Control, 7, 136-141. Zhang S.J., Yu H.Q., (2004), Radiation-induced degradation of polyvinyl alcohol in aqueous solutions, Water Research, 38, 309-316. Zhang S.J., Yu H.Q., Ge X.W., Zhu R.F., (2005), Optimization of radiolytic degradation of poly(vinyl alcohol), Industrial & Engineering Chemistry Research, 44, 1995-2001. Zhang Y., Gu R., (2007), Treatment of PVA-size wastewater with a ceramic dynamic membrane. Journal of Donghua University, 33, 236-238.


July 2013, Vol.12, No. 7, 1329-1334



EFFECT OF AERATION RATE ON DOMESTIC WASTEWATER TREATMENT USING AN INTERMITTENTLY AERATED SEQUENCING BATCH REACTOR (IASBR) TECHNOLOGY Liam G. Henry1, Liwen Xiao2, Xinmin Zhan1 1 Civil Engineering, National University of Ireland, Galway, Ireland Department of Civil, Structural and Environmental Engineering, Trinity College Dublin, Dublin 2, Ireland

2

Abstract Effects of the aeration rate on nutrient removal from synthetic domestic wastewater using an intermittently aerated sequencing batch reactor (IASBR) were studied at ambient temperature. Two aeration rates, 0.8 and 1 L air/min, were studied. At the aeration rate of 0.8 L air/min, removals of COD, NH4+-N, total nitrogen (TN) and PO43--P were 84%, 96%, 75% and 99%, respectively. At the aeration rate of 1 L air/min, removals of COD, NH4+-N, TN and PO43--P were 90%, 99%, 70% and 66%, respectively. The increased DO concentrations resulted in improved nitrification, but interfered with denitrification and enhanced biological phosphorus removal (EBPR). Key words: aeration, denitrification, EBPR, IASBR, nitrification Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Nutrient enrichment in freshwater, estuarine and marine water bodies is a significant environmental issue worldwide. A water quality survey taken in the period of 2007 – 2009 shows that in Ireland, 27 river sites are classified as seriously polluted, 11.3% of lakes are classified as less than satisfactory and 15.3% of groundwater is classified at a poor status, with the primary pollutant within the groundwater being phosphate (Irish EPA, 2011). To achieve the target set by the European Commission Water Framework Directive (EC, 2000) that all water bodies in the EU have to achieve a ‘good status’ by 2015, Ireland is under pressure to treat municipal wastewater to high standards prior to discharge into the environment. A report on urban wastewater discharges in Ireland (Irish EPA, 2012) has found that over half (58%) of wastewater treatment plants have failed to achieve the quality set by the standards and 

guidelines of the Urban Waste Water Treatment Regulations 2001-2010 and the 1991 Urban Waste Water Treatment Directive (UWWTD). Biological processes utilizing one or more unaerated and aerated zones have been widely used to remove carbon, nitrogen (N) and phosphorus (P) from wastewater. While biological carbon removal can occur in both aerobic and anaerobic conditions, biological nitrogen removal requires alternating aerobic and anoxic conditions and biological phosphorus removal requires alternating anaerobic and aerobic conditions (Mehrali et al., 2012; Wu et al., 2011). To achieve alternating aerobic, anaerobic and anoxic conditions for simultaneous N and P removal, intermittently aerated sequencing batch reactor (IASBR) technology was developed. It combines the benefits of modest capital investment and operational expenditure and has been successfully used to treat a range of wastewaters such as domestic wastewater (Zhao et al., 1999), swine wastewater (Mota et al.,

Author to whom all correspondence should be addressed: E-mail: [email protected]; Phone: +353 91495239

Henry et al./Environmental Engineering and Management Journal 12 (2013), 7, 1329-1334

2005), slaughterhouse wastewater (Li et al., 2008a, 2008b), textile industry wastewater (Giorgetti et al., 2011) and wastewater generated in meat product companies (Rodriguez et al., 2011). In their study, Li et al. (2011) utilized IASBRs to achieve efficient partial nitrification at moderately low temperature, which would save the aeration cost and have the capacity to treat a wide range of ammonium-rich wastewaters with low COD:N ratios. Zhang et al. (2011) investigated the treatment of separated digestate liquid after anaerobic digestion of pig manure in an IASBR and achieved COD and total nitrogen (TN) removals of 89.8% and 76.5%, respectively. These studies display the flexibility of IASBRs in the removal of COD and nutrients from wastewater. The aeration rate is an important operational parameter to the IASBR technology. If poorly controlled it can negatively impact the performance and result in increased running costs. In this study the performance of a laboratory scale IASBR in domestic wastewater treatment at two aeration rates - 0.8 and 1 L air/min - was studied and compared. 2. Materials and methods 2.1. Materials An IASBR unit was made of stainless steel, with a working volume of 5 liters, and was operated at ambient temperature. The IASBR was automated, with pumps, stirrers and air pumps controlled by a Siemens Programmable Logic Controller (PLC) using Step 7 MicroWIN software. The operation sequence of the IASBR is detailed in Fig. 1, with a sequencing batch reactor (SBR) cycle duration of 6 hours. The IASBR had four cycles per day; in each cycle, there were three alternating 50-minute non-aeration and 50-minute aeration periods. During each cycle 3 liters of synthetic wastewater was treated, with two Masterflex L/S peristaltic pumps used to fill wastewater and withdraw treated wastewater. The reactor was constantly stirred with a magnetic stirrer during the fill, non-aeration and aeration periods. During the aeration periods air was supplied using an aquarium air pump through permeable stone

diffusers located at the bottom of the reactor; the air flow rate was regulated by an air flow meter. Once a day, 500 mL of mixed liquor was withdrawn from the reactor just before the settle phase, resulting in a sludge retention time (SRT) of 10 days (without consideration of solids loss in the treated wastewater). Dissolved oxygen (DO), pH and oxidation reduction potential (ORP) were real time monitored using electrodes (HI-9828 multi-parameter electrode, Hanna Instruments, United Kingdom). 2.2. Synthetic wastewater and seed sludge The components of the synthetic wastewater included: 403 mg/L sodium acetate, 30 mg/L yeast extract, 120 mg/L dried milk, 30 mg/L urea, 60 mg/L NH4Cl, 100 mg/L Na2HPO4.12H2O, 50 mg/L KHCO3, 130 mg/L NaHCO3, 50 mg/L MgSO4.7H2O, 2 mg/L FeSO4.7H2O, 2 mg/L MnSO4.H2O, and 3 mg/L CaCl2.6H2O. Synthetic wastewater was made every 1 – 2 days, stored beside the IASBR unit and stirred by 2 magnetic stirrers constantly. The average concentrations of total chemical oxygen demand (CODt), biological oxygen demand (BOD5), total nitrogen (TN) and total phosphorus (TP) were 346 mg/L, 230 mg/L, 33 mg/L and 10 mg/L, respectively. The use of synthetic wastewater allowed concentrations to remain consistent in the influent during the study, with little fluctuation. The reactor was seeded with activated sludge taken from the return activated sludge (RAS) line at the municipal wastewater treatment plant in Athenry, Co. Galway. The IASBR unit was operated at a low aeration rate (LAR) of 0.8 L air/min during Days 1 – 22 and a high aeration rate (HAR) of 1 L air/min during Days 23 - 55. At the LAR the specific organic loading rate (SOLR) inside the IASBR was 0.24 g BOD/g MLSS.d. At the increased aeration rate the SOLR was 0.2 g BOD/g MLSS.d. 2.3. Analytical methods Ammonium (NH4+-N), nitrite (NO2--N), nitrate (NO3--N) and ortho-phosphate (PO43--P) were analyzed using a Konelab 20 analyzer (Thermo Clinical Labsystems, Vantaa, Finland).

Fig. 1. IASBR operation sequences

1330

Effect of aeration rate on treatment of domestic wastewater using an intermittently aerated Sequencing Batch Reactor (IASBR)

Suspended solids (SS) and mixed liquor suspended solids (MLSS) were measured in accordance with the standard APHA methods (APHA, 1995). The pore size of the filter paper for SS and MLSS measurements was 0.45 µm. BOD5 concentrations were measured with an OxiDirect BOD-System (Lovibond, UK). COD concentrations were measured using the closed reflux method and a Hach spectrophotometer (model DR/2010; Hach, USA). TN and TP were measured using TN and TP kits according to the manufacturers’ protocol (Hach, USA). Polyhydroxybutyrate (PHB) contents in activated sludge were measured using the procedure adopted by Rodgers et al. (2010). High performance liquid chromatography (HPLC, Agilent 1200, Agilent Technology, USA) was used to detect PHB with a UV index detector and an Animex HPX-87H column (Bio-Rad, USA). Separation during HPLC tests was achieved using a mobile phase of 1% (vol/vol) H2SO4 at a flow rate of 0.1 mL/min, a column temperature of 65 oC, and a refractive index detector temperature of 40 oC. 3. Results and discussion 3.1. Effects of the aeration rate on the overall performance of the IASBR unit The performance of the IASBR at the two aeration rates is given in Table 1. At the LAR the nutrient concentrations in the effluent were within the emission standards set by the Irish EPA. However at the HAR, the system failed to reach the emission standards for phosphorus, with the average concentration of 3.4 mg PO43--P/L and a peak concentration of 7.0 mg PO43--P/L in the effluent. The HAR deteriorated the P removal performance of the system even though it improved COD, BOD5 and NH4+-N removal. NH4+-N concentrations in the IASBR effluent decreased from 0.9 mg/L to 0.4 mg/L when the aeration rate was increased from 0.8 L air/min to 1 L air/min. However, the HAR deteriorated the TN removal performance of the reactor. The effluent TN concentration rose from 8.5 mg/L at the LAR to 10.2 mg/L at the HAR. MLSS concentrations varied slightly at each aeration rate. At the LAR the average MLSS concentration was 2.3 g/L. At the HAR the concentration increased slightly to 2.7 g/L. At the LAR, the average sludge volume index (SVI) value was 204 mL/g but it was 113 mL/g at the HAR, on average, indicating that the sludge settling property was improved at the HAR. The reduced SVI at the HAR resulted in reduced effluent SS concentrations. Yang et al. (2010) used a novel sequencing batch moving bed membrane bioreactor (SBMBMBR) for simultaneous nitrogen and phosphorus removal from wastewater, and achieved removal efficiencies of 93.5%, 82.6%, 95.6% and 84.1% for COD, TN, NH4+-N and TP, respectively.

These results are similar to those observed in this study (Table 1). They also observed a rapid decrease in phosphorus removal as the length of the aeration period was increased and that high DO concentration in the aeration period (up to 6 mg/L) inhibited the N and P removal. Their research confirms that the operation of an SBR would be highly affected by the aeration rate. 3.2. Phase study at the two aeration rates At the LAR the majority of NH4+-N removal occurred in the aeration periods, with the reduction of 5, 6.5 and 2.2 mg NH4+-N/L in aeration periods 1, 2 and 3, respectively (Fig. 2a). DO concentrations peaked at 0.6, 1.6 and 5.1 mg/L at the end of aeration periods 1, 2 and 3, respectively (Fig. 2b). When the aeration ceased, the oxygen concentrations quickly dropped to zero. In aeration period 3 the DO jumped during the first 5-10 minutes of the aeration period, and then increased very slowly and leveled off, before it quickly increased. The first DO jump could be due to the delay of the nitrification process; this phenomenon has been observed by Zhan et al. (2009) and Gieseke et al. (2002). When the nitrification process occurred, oxygen was consumed and DO concentrations leveled off. After completion of the nitrification process, DO concentrations started to increase again. pH was between 7.3 and 8. The peak value was observed in the middle of the non-aeration period (Minute 120; Fig. 2b). It is well known that the nitrification process decreases the pH value and denitrification increases pH. During aeration period 3 the pH leveled off after a reduction and then began to increase. This point, referred to as the pH valley, coincided with the rapid increase of DO indicating completion of the nitrification process. At the HAR the majority of NH4+-N removal occurred in aeration period 1 with the reduction of 11 mg NH4+-N/L (Fig. 3a). 1.5 mg NH4+-N/L and 2.5 mg NH4+-N/L were removed in non-aeration period 2 and aeration period 2, respectively. The DO concentrations peaked at 3 mg/L, 9 mg/L and 9 mg/L, respectively, at the end of aeration periods 1, 2 and 3 (Fig. 3b). It slowly dropped to zero during nonaeration period 3 after the aeration was terminated. The rapid increase of DO concentrations prevented a clear observation of the DO control point. The pH gradually decreased from 8.7 after the fill phase to about 7.8 at the end of aeration period 1 and then fluctuated between 7.7 and 8.2 in the rest of the cycle. The ‘pH valley’ was observed during aeration period 2; however it occurred so rapidly that it would be difficult to be used as an operation control point. At the LAR the total oxidized nitrogen (TON) (NO3--N plus NO2--N) production in aeration periods was 13.6 mg TON/L, with the production of 3.1, 8.0 and 2.6 mg TON/L in aeration periods 1, 2 and 3, respectively, indicating that NH4+-N removed during the aeration periods was completely converted to NO3--N and NO2--N.

1331


Table 1. Performance of IASBR at 0.8 and 1 L air/min aeration rates Parameter

Emission Standards

SS BOD5 COD TN TP NH4+-N PO43--P

60 20 125-250 or >75% removal 15-40 or >80% removal 10 -

IASBR Effluent (mg/L)* LAR HAR 13.6 (2.6) 10.6 (6.0) 4.5 (1.7) 4 (1.4) 53.6 (14.7) 38 (2.8) 8.5 (2.9) 10.2 (1.5) 1.3 (0.3) 5 (-) 0.9 (1.6) 0.4 (0.13) 0.07 (0.06) 3.4 (2.2)

Removal Rate (%) LAR HAR 98 98.3 84 89.8 74 69 87 50 95.9 98.1 99.2 61.5

*Numbers in bracket are standard deviation

a)

b)

Fig. 2. NH4+-N, NO2--N, PO43--P and NO3--N concentrations: a) and DO, pH and ORP; b) in a typical cycle at the LAR

a)

b)

Fig. 3. NH4+-N, NO2--N, PO43--P and NO3--N concentrations: a) and DO, pH and ORP; b) in a typical cycle at the HAR

The denitrification process reduced the TON during non-aeration periods 2 and 3. At the HAR the TON production mainly occurred in aeration period 1, with the production of 10.5 mg NO3--N/L and 1.7 mg NO2--N /L. NH4+-N removed in aeration period 1 was completely converted to NO3--N and NO2--N. Denitrification was limited during the HAR. This was due to DO remaining in the non-aeration periods following the aeration periods.

1332

At the LAR, P concentrations decreased from 54 mg/L at the start of aeration period 1 to 14, 10, 2.0 and 1.0 mg/L at the end of aeration period 1, nonaeration period 2, aeration period 2 and aeration period 3, respectively. At the HAR, the high DO concentrations resulted in the collapse of the EBPR process (Brdjanovic et al., 1998). P concentrations decreased from 35 mg/L at the start of aeration period 1 to 10, 8.5 and 6 mg/L at the end of aeration

Effect of aeration rate on treatment of domestic wastewater using an intermittently aerated Sequencing Batch Reactor (IASBR)

period 1, non-aeration period 2 and aeration period 3, respectively. Zhu et al. (2011) compared the effectiveness of a conventional SBR (cSBR) to a humus soil SBR (HS-SBR) for the treatment of phosphorus. PO43--P removals were observed to be 97% and 80% in the HS-SBR and cSBR, respectively, less than the P removal observed at the LAR in this study. The peak and average uptake rates of PO43--P at the LAR were 1.54 and 0.63 mmol P/hr (or 47.8 and 19.6 mg P/l hr), respectively. The overall uptake rates observed by Zhu et al. (2011) were 1.98 mmol P/hr and 1.45 mmol P/hr in the HSSBR and cSBR, respectively. Kargi et al. (2005) tested PO43--P release and uptake rates with different carbon sources and observed a peak PO43--P uptake rate of 8.1 mg P/L.hr. The high PO43--P uptake levels confirm the occurrence of EBPR in the IASBR at the LAR. Fig. 4 shows the variation of the PHB content in activated sludge and PO43--P concentrations in a typical operation cycle at LAR. It was observed that during the first non-aeration period, as the PO43--P concentration rose, the PHB content in the activated sludge rose.

Fig. 4. Profiles of PHB content in activated sludge and PO43--P concentrations in a typical cycle at LAR

The PHB content during non-aeration periods 2 and 3 experienced a reduction. During these periods a decrease in the PHB content was not expected due to the anaerobic conditions. This decrease was accompanied by small PO43--P uptake which occurred during these periods. 4. Conclusions This study examined the effects of two aeration rates on the performance of an IASBR system. The following results were obtained: 1. The 0.8 L air/min aeration rate was considered better. Removals of COD, TN and TP were up to 84%, 78% and 90%, respectively. 2. An increased aeration rate of 1 L air/min improved nitrification. P removal declined due to the increased DO concentrations in the react period affecting P uptake and release. Removals of COD, TN and TP were up to 90%, 70% and 35%, respectively.

This study clearly shows that it is important to choose a proper aeration rate for a good performance of IASBRs. Acknowledgements The authors acknowledge the financial support of the EPA (Ref.: 2009-PhD-ET-7). They would also like to thank the technicians and Dr. Mark Healy in the Civil Engineering Department of NUI Galway.

References Brdjanovic D., Slamet A., Van Loosdrecht M.C.M., Alaerts G.J., Heijnen J.J., (1998), Impact of excessive aeration on biological phosphorus removal from wastewater, Water Research, 32, 200-208. Gieseke A., Arnz P., Amann R., Schramm A., (2002), Simultaneous P and N removal in a sequencing batch biofilm reactor: insights from reactor and microscale investigations, Water Research, 36, 501-509 Giorgetti L., Talouizte H., Merzouki M., Caltavuturo L., Geri C., Frassinetti S., (2011), Genotoxicity evaluation of effluents from textile industries of the region FezBoulmane, Morocco: A Case Study, Ecotoxicology and Environmental Safety, 74, 2275-2283. Irish EPA., (2006), Draft BAT guidance note on best available techniques for the slaughtering sector, Irish EPA, Wexford. Irish EPA., (2011), The provision and quality of drinking water in Ireland, Report for the years 2008 – 2009, Irish EPA, Wexford. Irish EPA., (2012), Focus on Urban Waste Water Discharges in Ireland, Report for the year 2009, Irish EPA, Wexford. Kargi F., Uygur A., Savas Baskaya, H., (2005), Phosphate uptake and release rates with different carbon sources in biological nutrient removal using a sequencing batch reactor, Journal of Environmental Management, 76, 71–75. Li J.P., Healy M.G., Zhan X.M., Rodgers M., (2008a), Nutrient removal from slaughterhouse wastewater in an intermittently aerated sequencing batch reactor, Bioresource Technology, 99, 7644–7650. Li J.P., Healy M.G., Zhan X.M., Norton D., Rodgers M., (2008b), Effect of aeration rate on nutrient removal from slaughterhouse wastewater in intermittently aerated sequencing batch reactors, Water Air and Soil Pollution, 192, 251–261. Li J.P., Elliott D., Nielsen M., Healy M.G., Zhan X.M., (2011), Long-term partial nitrification in an intermittently aerated sequencing batch reactor (SBR) treating ammonium-rich wastewater under controlled oxygen-limited condition, Biochemical Engineering Journal, 55, 215-222. Mehrali S., Moghaddam M.R.A., Hashemi S.H., (2012), Feasibility study of several cyclic anaerobic/aerobic conditions in sbr system for treating of simulated dye (Reactive Blue 19) wastewater, Environmental Engineering and Management Journal, 11, 617-621. Mota C., Head M., Ridenoure J.A., Cheng J.J, De los Reyes F.L., (2005), Effects of aeration cycles on nitrifying bacterial populations and nitrogen removal in intermittently aerated reactor, Applied and Environmental Microbiology, 71, 8565-8572. Rodgers M., Wu G., (2010), Production of polyhydroxybutyrate by activated sludge performing enhanced biological phosphorus removal, Bioresource Technology, 101, 1049-1053.

1333


Rodriguez D.C., Pino N., Penuela G., (2011), Monitoring the removal of nitrogen by applying a nitrificationdenitrification process in a sequencing batch reactor (SBR), Bioresource Technology, 102, 2316-2321. Wu G.X., Guan Y.T., Zhan X.M., (2011), Characteristics of nutrient removal from synthetic wastewater with different organic substrates, Environmental Engineering and Management Journal, 10, 649-654.. Yang S., Yang F., Fu Z., Wang T., Lei R., (2010), Simultaneous nitrogen and phosphorus removal by a novel sequencing batch moving bed membrane bioreactor for wastewater treatment, Journal of Hazardous Materials, 175, 551-557. Zhan X., Healy M.G., Li J., (2009), Nitrogen removal from slaughterhouse wastewater in a sequencing batch reactor under controlled low DO conditions, Bioprocess Biosystem Engineering, 32, 607-614.

1334

Zhang M., Lawlor P.G., Wu G., Lynch B., Zhan X. (2011), Partial nitrification, and nutrient removal in intermittently aerated sequencing batch reactors treating separated digestate liquid after non-aeration digestion of pig manure, Bioprocess Biosystem Engineering, 34, 1049-1056. Zhao H.W., Mavinic D.S., Oldham W.K., Koch F.A., (1999), Controlling factors for simultaneous nitrification and denitrification in a two-stage intermittent aeration process treating domestic sewage, Water Research, 33, 961-970. Zhu R., Wu M., Zhu H., Wang Y., Yang J., (2011), Enhanced phosphorus removal by a humus soil cooperated sequencing batch reactor, using acetate as carbon source, Chemical Engineering Journal, 166, 687-692.


July 2013, Vol.12, No. 7, 1335-1344



CHARACTERIZATION OF NITRIFICATION PERFORMANCE AND MICROBIAL COMMUNITY IN A MBBR AND INTEGRATED GBBR-MBBR TREATING HEAVILY POLLUTED RIVER WATER Xiangchun Quan, Linyun Gu, Yin Qian, Yuansheng Pei, Zhifeng Yang State Key Joint Laboratory of Environmental Simulation and Pollution Control/Key Laboratory of Water and Sediment Sciences of Ministry of Education, School of Environment, Beijing Normal University, Beijing 100875, P. R. China

Abstract A one-stage aerobic moving bed biofilm reactor (MBBR) (Reactor A) and a combined reactor (Reactor B) involving an anoxic gravel-bed biofilm reactor (GBBR) and an aerobic MBBR were applied to the treatment of heavily polluted river water. Reactor performance was investigated throughout the experiment for almost 200 days and molecular techniques including PCR-DGGE, FISH/CLSM and FISH/FCM were used to reveal the evolutions of bacteria community, abundance of nitrifying bacteria and their spatial distribution in biofilms for the comparative study of the two reactors. Results show that Reactor B performed better than Reactor A in pollutants removal with COD, ammonia and TN removals enhanced by 6-16%, 32-59%, and 9-31%, respectively. In addition, Reactor B was more stable towards the increasing of organic and ammonia loadings. The aerobic biofilms in Reactor B were thinner and occupied by large nitrifying populations (16-41% of the total bacteria) than the corresponding part in Reactor A (9-22% of the total bacteria). On the whole, the integrated GBBR-MBBR was more efficient and lower-cost compared to the one-stage MBBR and therefore more suitable for the treatment of the river water. Key words: ammonia oxidation, biofilm, microbial community, moving bed biofilm reactor, nitrifying bacteria Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Water bodies including rivers are heavily polluted today in China due to the discharge of large quantities of wastewater at non-point sources. After treatment, wastewater could hardly meet the national standards for discharge, in most cases due to process inefficiency and operation problems in wastewater treatment plants. These wastewaters generally contain a high content of organic compounds and nitrogen. Remediation of rivers receiving large quantities of such wastewater is a challenging task. A variety of technologies have been applied to treat polluted rivers, such as chemical oxidation (Goncharuk et al., 2009; Zaharia et al., 2011), artificial wetlands and enhanced biodegradation 

(Ruan et al., 2006; Juang et al., 2008). Among these, chemical oxidation has been rarely used in practice due to its high costs. Application of artificial wetlands is also restricted due to large area of land requirement. Biotechnology, especially biofilm based biological method, is now more attractive in river remediation for the advantages of effectiveness, low investment and operation cost (Nava-Arenas et al., 2012). Biofilms are generally characterized by high cell density and metabolic activity. They can be easily formed on a variety of natural and artificial carriers/substrates like sands, gravels and plastic rings. Although many different types of biofilm technologies have been developed for wastewater treatment, such as biological aerated filters (Rother et al., 2002), fluidized bed reactors (Pruden et al.,

Author to whom all correspondence should be addressed: E-mail: [email protected]; Phone/Fax: 86-10-58802374

Quan et al./Environmental Engineering and Management Journal 12 (2013), 7, 1335-1344

2005), suspended carrier biofilm reactors (Wang et al., 2005) and biological oxidation contactors (Wyffels et al., 2003), these technologies can hardly been applied to natural river remediation directly due to their complexity of the physical environment. A cost-effective process based on biofilm approach, that is suitable for river remediation, deserves further study. Recently, moving bed biofilm reactors (MBBR) have become a promising technology for the removal of organic matters and nutrients (N and P) (Hem et al., 1994; Luostarinen et al., 2006). In a MBBR, bacteria are fixed in biofilms on carriers, which are suspended and can move freely in the reactor. Therefore, MBBR provides a long retention time for the biomass and accommodate high loading rates without the problem of clogging. Although MBBR has been widely used for posttreatment of wastewater, it has rarely been applied in the purification of polluted rivers. Ammonia removal is generally an important consideration for process selection and design, because nitrifying bacteria grow slowly and are very sensitive to many external factors like carbon sources, oxygen concentrations, pH, and temperature changes (Haseborg et al., 2010; Xia et al., 2010). Nitrifying bacteria, as autotrophic microorganisms, may be suppressed by the increase of organic carbon loading for the fast growing of heterotrophic microorganisms in the presence of high concentrations of organic compounds (Akker et al., 2011; Pearce and Edwards 2010). The competition for dissolved oxygen (DO) between heterotrophic bacteria and nitrifying bacteria may also restrain ammonia oxidation when oxygen becomes limitation (Nogueira et al., 2002; Terada et al., 2003). In order to reveal the relationship between the nitrifying performance and the microbes (such as ammonia-oxidizing bacteria (AOB) and nitriteoxidizing bacteria (NOB)) involved in the nitrifying process, the Denaturing Gradient Gel Electrophoresis (DGGE) was conducted to reveal the dynamics of bacteria community and the Fluorescent In Situ Hybridization (FISH) method coupled with Confocal Laser Scanning Microscopy (CLSM) was used to characterize the abundance of nitrifying bacteria and their spatial distribution in biofilms In this study, two biofilm processes, an aerobic MBBR and a reactor comprising an aerobic MBBR and an anoxic gravel-bed biofilm reactor (GBBR) which provided pretreatment, were used to treat synthetic polluted river water. This study aimed to characterize the performance of the two biofilm reactors in pollutants removal especially ammonia removal, and to investigate characteristics of the biofilm structure, microbial community, distribution and abundance of nitrifying microbial populations in biofilms. It is expected that this study will provide a better understanding of the performance of a biofilm process for high-strength ammonia polluted river water remediation. It is also expected that the study will provide valuable information for process design.

1336

2. Materials and methods 2.1. Experimental set-up and operation A schematic diagram of the MBBR (Reactor A) and integrated GBBR-MBBR (Reactor B) is shown in Fig. 1. Each reactor had a size of 80 cm×31 cm×35 cm and an effective volume of 72 L. Reactor A consisted of an aerobic reaction zone (48 L) filled with combined carriers and a settling zone (24 L). Reactor B had a similar structure as Reactor A except that the reaction zone was equally divided into an anoxic zone and an aerobic zone (24L each). The anoxic zone was not aerated and filled with gravels and volcanic rocks as the biofilm carriers. The aerobic zone was packed with the same carriers as Reactor A. The combined carriers were made by adding small suspended and cylindrical carriers (1.8 cm in length and 1.85 cm in diameter) to lattice structured plastic balls (8.8 cm in diameter). The small suspended carriers can move up and down in the balls and along with the ball. This structure allowed easy application and retaining the suspended carriers in reactors. The two reactors were firstly inoculated with activated sludge taken from a local municipal wastewater treatment plant (Gaobeidian wastewater treatment plant, Beijing, China) and then fed with synthetic polluted river water. The influent chemical oxygen demand (COD), ammonium nitrogen (NH4+N) and total nitrogen (TN) concentrations increased gradually throughout the operation period. The whole period can be divided into four stages: Start-up (Days 1 to 21), Stage I (Days 22 to 85), Stage II (Days 86 to 144) and Stage III (Days 145 to 184). The synthetic water contained the following compositions: glucose 141-234 mg/L, starch 61-92 mg/L, NH4Cl 69-138 mg/L, KNO3 15 mg/L, Na2HPO4•12H2O 26.7 mg/L, KH2PO4•2H2O 15.2 mg/L, CaCl2•2H2O 10 mg/L, MgSO4 8 mg/L, peptone 10 mg/L, and humic acid 10 mg/L. Both reactors had a hydraulic retention time (HRT) of 8 hrs for the reaction zone. Air bubbles were supplied to the reactors through dispensers installed at the bottom of aerobic reaction zones at a flow rate of 4-6 L/min for Reactor A and 1.5-3 L/min for the aerobic zone of Reactor B with the oxygen concentration controlled at 5-8 mg/L. The anoxic zone of Reactor B was not aerated and the DO was generally below 1.0 mg/L. Water temperature varied slightly in the range of 18-25 °C. Some biofilm carriers were sampled from the two reactors when steady state was reached to determine their nitrification potentials. 2.2. DNA extraction and PCR-DGGE analysis Genomic DNA was extracted from the boifilm samples withdrawn at different operation times using the EZ-10 Spin Column Bacterial Genomic DNA MiniPreps Kits (Bio Basic Inc, Canada).

Characterization of nitrification performance and microbial community in a MBBR and integrated GBBR-MBBR

Fig. 1. A schematic diagram of the MBBR (Reactor A) and integrated GBBR-MBBR (Reactor B) (left) and structure of the combined carriers (right)

Bacterial 16S rDNA fragments of the biofilm samples were amplified by Polymerase Chain Reaction (PCR) using the primers 341F-GC and 907 R (Muyzer et al., 1993; Teske et al., 1996). One PCR reaction (50 µL) contained: TaqDNA-polymerase (5 U/µL), 0.25 µL; GC buffer I (Mg2+ Plus), 25 µL; dNTP mixture (each 2.5 mM), 4 µL; DNA template, 1 µL; primer 341F-GC (10 μM), 1 µL; primer 907 R (10 μM), 1 µL; and sterilized MilliQ water, 18.75 µL. Amplification was performed with touchdown PCR under the following conditions: an initial denaturing step at 94 °C for 5 min; then 8 cycles of denaturing at 94 °C for 30 s, annealing at 63-56 °C for 1 min (decreasing by 1 °C each cycle) and extension at 72 °C for 90 s; followed by 25 cycles of 94 °C for 30 s, 56 °C for 1 min, 72 °C for 90 s; final extension at 72 °C for 7 min and then kept at 4 °C. The Denaturing Gradient Gel Electrophoresis (DGGE) was performed using a DCode universal mutation detection system (Bio-Rad, USA). PCR products (20 µL) were run on 6% acrylamide gels with a denaturing gradient of 35-55%. Electrophoresis was performed at 120 V for 12 hrs at 60 °C. Gels were stained with SYBR Green I and photographed using a GEL imaging system (VILBER INFINITY 3000, France). Bands of interest were excised from the gels and DNA was recovered from the target bands for sequence determination. The isolated sequences were compared with 16S rDNA sequences obtained via BLAST searches of the National Center for Biotechnology Information database (http://blast.ncbi.nlm.nih.gov). The scanned gels containing DNA band profiles were analyzed using quantity One 1-D analysis software. The Shannon-Weaver index of species diversity (H) was calculated to evaluate the microbial diversity according to the following Eq. (1):

s

H 

p

i

log( p i )

(1)

i 1

where pi is the proportion of band i in the DGGE profile and s is the total number of the bands. 2.3. FISH-CLSM/FCM analysis The Fluorescent in Situ Hybridization (FISH) method coupled with Confocal Laser Scanning Microscopy (CLSM) was used to determine the nitrifying bacteria distribution in the biofilms. Carrier samples with well grown biofilms were withdrawn from the aerobic zones of the two reactors at steady operation state. Biofilms were detached from carriers by ultrasonic and then washed with PBS for three times. Samples were first fixed with 4% paraformaldehyde for 3 h at 4°C, dehydrated using graded ethanol (50%, 80% and 96%), dried in the air, and then stocked at -20 °C. As the penetrability of oligonucleotide probes and the transmittance of light would be inhibited when the thickness of biofilm reached more than 100 μm (Barranguet et al., 2004), biofilms in current study were cut into thin slices of 50 μm in thickness by a freezing microtome (Leica, German) at -20°C. The slices were fixed on the glass slides. 2 μL probes at a concentration of 100 ng/μL and 8 μL hybridization buffer was added to the samples drop by drop and hybridized in darkness at 46°C for 3 h, and then the slides were dipped into a washing solution at 48 °C three times each for 3 min. After being rinsed in distilled water once, the slides were dried in the air for further analyses. The oligonucleotide probes used in the FISH experiments (Egli et al., 2003; Coskuner et al., 2005) included EUB 338 targeting almost all bacteria, Nso 190

1337

Quan et al./Environmental Engineering and Management Journal 12 (2013), 7, 1335-1344

targeting ammonia-oxidizing Betaproteobacteria and Nit3 targeting Nitrobacter. The finally obtained samples were observed under a CLSM microscopy (Carl Zeiss, German). FISH method coupled with Flow Cytometry (FCM) method was used to measure the fractions of AOB and NOB in biofilms. The biofilm samples pretreated with above procedures were added to centrifuge tubes, dispersed by an ultrasonic cell disruptor (Sonics, USA) and diluted into the concentration about 108 cell /mL. 100 μL of the diluted cell solution was concentrated by centrifugation, and the collected cells were added with 5 μL probes at a concentration of 100 ng/μL and 185 μL hybridization buffer drop by drop. The hybridization was controlled at 46°C for 3 h in darkness, and then the samples were eluted at 48 °C for at least 20 min. Finally, the hybridized samples were fully suspended for further FCM (BD, USA) examination. 2.4 Analytical methods The effluent water samples withdrawn from the reactors were filtered with 0.22 μm filter prior to determining the concentrations of COD and nitrogen. Mixed liquor suspended solids (MLSS), COD, NH4+N, TN, nitrite nitrogen (NO2-N) and nitrate nitrogen (NO3-N) were analyzed in accordance to the standard methods (APHA, 1998). DO and water temperature were monitored by Sens-ion6 portable DO meter (HACH, USA). 3. Results and discussion 3.1. Reactor performances The removal efficiencies for COD, ammonia and TN in the two reactors during the whole

operation period are presented in Table 1. Both reactors functioned successfully within 21 days with an average COD removal of 70.7% and 89.3% for Reactor A and Reactor B, respectively. When the influent COD concentrations gradually increased from 155 to 350 mg/L between Stage I to Stage II, the removal of COD was relatively stable with an average of 82.2% and 91.6% for Reactor A and Reactor B, respectively. The GBBR-MBBR combined reactor (Reactor B) showed a higher COD removal than MBBR alone (Reactor A), in which most of the influent COD (7883%) was removed in the anoxic GBBR and only 813% was removed in the followed aerobic MBBR (Fig. 2a). The GBBR-MBBR system also displayed higher ammonia removal than the MBBR with average removals attained 80.3% and 47.9% during Stage I, respectively. With the increase in COD and ammonia loadings from Stage I to Stage III, the average ammonia removal in Reactor A declined to 29.1%, while that in Reactor B remained above 80%, indicating that Reactor B was more stable than Reactor A in ammonia removal towards the increase of organic loadings. The decrease of ammonia removal with the increase of organic loadings was also reported in other studies (Jonoud et al., 2003). Most recently, Akker et al., (2011) found that increase in BOD adversely impacted the nitrification capacity of a trickling filter. This may be attributed to the stimulation of heterotrophic bacteria at high organic loadings, which would probably suppress the growth of autotrophic populations. Although the activity of nitrifying bacteria might also be inhibited by the free ammonia concentration in solution, its inhibition effect on nitrification could be ignored (Kim et al., 2006) as the ammonia concentrations in this study were below 100 mg/L.

Table 1. The influent and effluent water quality parameters at different operation stages in two reactors Operation Stages

Influent (mg/L) NH4+-N

CODCr

TN

A/B 235.7±22a 155.0±30.7

A/B 30.7±3.6 18.1±1.2

A/B 33.6±2.8 25.0±2.8

StageⅡ(86144d)

241.6±40.6

27.8±3.0

35.4±4.4

StageⅢ(145184d) Average

349.3±28.4

36.2±3.9

41.3±4.3

246.7

28.2

33.8

Start-up (1-21d) StageⅠ(2285d)

Operation Stages

CODCr A 69.15±27.4(70.7%)b 32.38±13.0 (78.2%) 41.62±13.5(82.6%)

B 25.19±25.2(89.3%) 9.34±9.3 (93.7%) 19.98±8.3 (91.4%)

Effluent (mg/L) CODCr A 21.08±3.3(30.0%) 9.45±3.5 (47.9%) 21.58±6.5(22.4%)

Start-up (1-21d) StageI (22-85d) StageII (86144d) StageIII (14548.98±11.4 (85.9%) 28.21±10.4(91.8%) 25.54±4.1(29.1%) 184d) Average 40.99 (82.2%) 19.18 (91.6%) 19.41 (31.9%) a Standard deviation and b Average removal in the brackets

1338

B 16.52±8.1(44.1%) 3.56±1.7 (80.3%) 5.23±2.6 (81.0%)

CODCr A B 24.26±4.7(26.9%) 17.28±3.0(48.1%) 16.75±2.4(32.6%) 14.46±2.4 (41.5%) 26.52±5.9(25.3%) 15.62±4.9 (56.1%)

4.33±2.5 (88.1%)

27.09±4.7(34.1%)

17.50±3.3 (57.3%)

7.41 (83.1%)

23.66 (30.7%)

16.21 (51.6%)

Characterization of nitrification performance and microbial community in a MBBR and integrated GBBR-MBBR

TN Conc. (mgL)

b

100 Influent 1

st

80

effluent

nd

60

2 effluent Total removal

40 20

COD Removal (%)

500 450 400 350 300 250 200 150 100 50 0 80

0 80

70 60

60 50

40

40 20

30 20

0

TN Removal (%)

COD Conc. (mg/L)

a

10 0

20

40

60

80

100

120

140

160

180

200

Fig. 2. COD (a) and TN (b) removal in the integrated GBBR-MBBR reactor (1st effluent represents the effluent from the anoxic GBBR and 2nd effluent represents the effluent from the aerobic MBBR)

For the combined reactor, as the GBBR took an important role in removing organic compounds, the nitrification efficiency was less influenced by the increase of organic loadings. The GBBR-MBBR system also demonstrated better performance in TN removal than the one-stage MBBR system, with TN removals ranged 25.332.6% and 41.5-57.3%, respectively. Increase in organic loadings from Stage I to Stage III led to an increase in TN removal from 41.5% to 57.3% in the Reactor B possibly. This is probably due to the increased carbon source available for denitrification, while TN removal in Reactor A was less influenced. The poor removal of TN in Reactor A may be attributed to the incomplete oxidation of ammonia. Biofilms withdrawn from the two reactors at the end of operation also demonstrated different nitrification rates. The combined system had nitrification rates of 1.1 mg NH4+-N/g MLSS·h-1 and 22.49 mg NH4+-N/g MLSS·h-1for the anoxic biofilms in the GBBR and the aerobic biofilms in the MBBR, respectively. Biofilms in Reactor A had a nitrification rate of 5.44 mg NH4+-N/g MLSS·h-1. These data further proved that biofilms in the integrated system displayed better nitrification potential than the single MBBR system.

From above comparisons between the two systems, it was found that GBBR-MBBR performed better than the one-stage MBBR in pollutants removal, with COD, ammonia and TN removal enhanced by 6-16%, 32-59%, and 9-31%, respectively. This could be explained by their different structures. For the combined reactor, because most of influent COD (about 80%) and a small fraction of ammonia (iron oxide scale>green ore>red ore. However, NH4+-N and TN show a removal hysteresis. The results support that the mechanism of phosphorous removal using dissimilatory Fe(III) reduction seems to be the chemical sedimentation driven by IRB, rather than the surface adsorption of Fe(III) source. Judging from ΔGθ, it is reasonable to believe that Feammox, ANAMMOX and nitrate-dependent oxidation coupled with Fe(III) reduction under anaerobic condition are thermodynamically feasible. They are potentially critical components of N cycle in activated sludge. Key words: activated sludge, dissimilatory Fe(III) reduction, nitrogen and phosphorus removal, reduction mechanism Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Microbial dissimilatory Fe(III) reduction, also known as ferric respiration, is a specific microbial enzymatic reaction. The iron-reducing bacteria (IRB) are the core of this reaction. It can be dated back in the 19th century that the phenomenon of Fe(III) reduction was discovered (Lovely and Phillips, 1988). Until recently, increased attention has been paid to the dissimilatory Fe(III) reduction and its application in water pollution control and environmental remediation. Research shows that microbial dissimilatory Fe(III) reduction reaction/process can promote oxidative degradation of organics (especially refractory organics), mineralize toxic heavy metal ions (such as uranium, chromium, arsenic etc.) and participate in the cycles of nitrogen and phosphorus in nature (Botton et al., 

2007; Li et al., 2010; Orbeci et al., 2012). Meanwhile, IRB is an important microbial group in wastewater treatment engineering. It accounts for about 3% of the total microbial population in activated sludge (Nielsen et al., 2002). Hence, research on strengthening the process of microbial dissimilatory Fe(III) reduction in the wastewater treatment system has a crucial importance for raising degradation efficiency of refractory organics, as well as improving the nitrogen and phosphorus removal. It is reasonable to believe that the different forms of Fe(III) may play different role in microbial dissimilatory Fe(III) reduction and their impacts on the activated sludge process, especially for nitrogen and phosphorus removal, remain unclear. Undoubtedly, seeking the cheap and easily microbial available Fe(III) seems very important to further develop the IRB-activated sludge process to achieve

Author to whom all correspondence should be addressed: E-mail: [email protected]; Phone: +86 931 4957057; Fax: +86 931 4956017

Wang et al./Environmental Engineering and Management Journal 12 (2013), 7, 1345-1352

The precipitate was washed by distilled water for six times to completely remove chloride and sodium (Lovley and Phillips, 1986). The iron concentration in the suspension of Fe(OH)3 was 5.188g/L. The iron oxide scale, with iron content of 71.3% (w/w), was supplied from an iron and steel plant in Beijing, China. The green ore and red ore, with iron content of 60.7% and 36.1% (w/w), respectively, were supplied from an iron ore plant in Henan, China. Their major mineral was hematite (Fe2O3). The oxide skin was prepared as 2.0~3.0mm strips while the ores were ground and sieved with the griddle of 80 mesh. Prepared Fe(III) materials were then used for the experiment. The surface morphology of different Fe(III) sources was observed by JSM-5600LV low vacuum scanning electron microscope with X-ray energy disperse spectroscopy (Japan Electronic Optical Company, American Kevex Company). As shown in Fig. 1, the iron oxide scale has dense structure, while iron ores show irregular and disperse.

better collaborative nitrogen and phosphorus removal. Ivanov et al., (2010) studied the economical efficiency of estrogens degradation in reject water of a MWWTP (municipal wastewater treatment plant) by respectively using iron ore and Fe(OH)3. Stabnikov et al., (2004) studied the effects of Fe (II) on phosphate removal, in which Fe(II) was derived from the wetland iron ore and iron-containing clay. The idea was that Fe(II) could react and precipitate with HPO42- in wastewater. Luan et al., (2009) investigated the reductive transformation of nitrobenzene by goethite, hematite, magnetite and steel converter slag bound Fe(III) system. It has been noted that more concerns and studies are focused on the application of pure iron-reducing bacteria in biodegradation of pollutants. Relatively, exploration of dissimilatory Fe(III) reduction process in activated sludge process and simultaneous nitrogen and phosphorus removal is rarely reported. In this study, cheap iron oxide scale, iron ore (such as green ore, red ore) and Fe(OH)3 were selected to be the electron acceptor of IRB, in which IRB were isolated and enrich-cultured from activated sludge in a SBBR (sequencing biofilm batch reactor) reactor. The dissimilatory Fe(III) reduction ability and its capacity to affect the removal of nitrogen and phosphate of urban sewage were investigated. In addition, the paper also aims to explore the microbial mediated multi-component interaction mechanisms, especially the C-Fe-N and C-Fe-P coupling mechanism, therefore, providing a better understanding of the dissimilatory Fe(III) reduction process.

2.1.2. Composition of seed sludge and synthetic wastewater The seed sludge was obtained from an SBBR reactor with an effective volume of 0.36m3. The reactor was filled with Fe0 and nano-attapulgite composite hydrophilicity polyurethane foam as filler and operated for wastewater treatment trial in the same laboratory for one year (Li et al., 2009). The seed activated sludge has preferable microbial activity. The ratio of mixed liquor suspended solid (MLSS) and mixed liquor volatile suspended solid (MLVSS) was 0.6, pH was about 7.2. The initial concentrations of Fe(II) and Fe(III) were 14.83 and 200.55mg/L, respectively. The inoculation amount of activated sludge was 1.41g MLSS/L. The simulated wastewater was used in the experiment and the formulation was as follows: 10mL glucose (10mmol/L); 5mL NH4Cl (5g/L); 1.5mL phosphate buffer (The molar ratio of KH2PO4 and K2HPO4 was 0.5809). No other trace elements were added. The initial content of TN, NH4+-N, TP in the simulated wastewater was 137.8, 121.0 and 21.8mg/L, respectively.

2. Materials and methods 2.1. Materials 2.1.1. Preparation of Fe(III) sources Iron oxide scale, green ore and red ore were employed as inexpensive sources of IRB. Their effectiveness are compared with Fe(OH)3, which is a common source used as IRB. A suspension of ferric hydroxide was prepared by slowly neutralizing 0.4mmol/L FeCl3 solution with 0.5 mol/L NaOH.

a)

b) Fig.1. Surface morphology of different Fe(III) materials under SEM observation: a) iron oxide scale; b) green ore; c) red ore

1346

c)

Effects of Fe(III) on dissimilatory ferric reduction, nitrogen and phosphorus removal in activated sludge process

2.2. Experimental setup The experiment was performed in strictly anaerobic batch cultivation in serum bottles with different Fe(III) sources. Before use, the serum bottles were autoclaved at 1210C for 20min. They were then put to a sterile room for cooling. 16.5mL simulated wastewater was poured into 100 mL serum bottles. Fe(OH)3 suspension with the content of iron 1.0 g/L was then added in one experimental serum bottle. Iron oxide scale and iron ores with the same quantity of iron was added in another experimental serum bottle. 20mL seed sludge, which was prepared by anaerobic treatment at 30±10C under seal for 7 days, was then introduced to the serum bottles with strictly aseptic manipulation. Two blank controls were performed following the same preparation of the experimental serum bottles by additional addition of 15% (v/v) ethanol to elimination microorganisms (Ivanov et al., 2010). Finally, distilled water was added to ensure a total volume of mixture 80mL. The pH of bottles was maintained 7.2±0.2. Each bottle was purged with nitrogen gas for 5 min and sealed up to ensure anaerobic conditions. Finally, the bottles were protected from light and kept with constant temperature of 30±10C for 25 days. 2.3. Sampling and chemical analysis Samples were taking in duplicate. One was used to determine the concentrations of TP, NH4+-N and TN and the other was used for total Fe(II) and Fe(III) measurement. TN, NH4+-N and TP were determined according to the standard methods (WWDA, 2002). To measure total Fe(II) and Fe(III), the samples were firstly acidified immediately after collection by the addition of 0.5mM HCl (in volume ration 1:4) for 30min to extract Fe(II) (Stabnikov et al., 2004). The samples were then placed in a centrifuge for 10min at 1500rpm. Thereafter, the centrifugate was filtered through 0.2 µm membrane

filter and the contents of total Fe(II) and Fe(III) was then determined via a modified phenanthroline method. All analytical determinations were performed in triplicate and mean values were reported. 3. Results and discussion 3.1. Effects of the IRB ability of different sources Fig. 1 illustrates the concentrations of total Fe(II) and Fe(III) during the cultivation process. Initial total Fe(II) concentration was 16.8±1.2mg/L for all samples. Concentration of Fe(II) remained unchanged in blank control, but increased over a period of 7 days in the experiments. It is noted that, at the period of 0 to 7 days, the dissociated Fe(II) of dissimilatory iron reduction in different Fe(III) sources exhibited approximately linear growth, indicating a well Fe(III) reducing ability. The computed maximum Fe(II) production rate at the iron source of Fe(OH)3, iron oxide scale, green ore and red ore was 27.58, 20.28, 16.88 and 15.27 mg/(L·d), respectively. Interestingly, the Fe(II) production rate had a statistically positive correlation with iron content of Fe(III) sources. However, after 7 days, the total Fe(II) concentration decreased rapidly and finally reach a stable level. It is worth noting that Fe(III) concentration of iron oxide scale system decreased gradually at 7 days, and then increased consistently even more significantly than Fe(OH)3 system after 16 days. It is probably due to the reason that iron oxide scale was a microbial available Fe(III) source. The striking drop of the total Fe(II) after 7 days may be contributed to the following reasons: 1). According to the literature (Zhu et al., 2012; Benner et al., 2002), magnetite could be produced at the process of microbial iron reduction, as shown in Eq. (1). This may cause the reduction of Fe(II); 2Fe(OH)+Fe2+→FeIIFe2IIIO4+2H++2H2O

(1)

200 180 160 140 120 100 80 60 40 20 0

0

2

4

6

8 10 12 14 16 18 20 22 24 26

Time/d Time (day)

TotalFe(III) Fe(III) concentration (mg/L) Total concentration/(mg/L)

Total Fe(II) concentration (mg/L) Total Fe(II) concentration/(mg/L)

220 500 450 400 350 300 250 200 150 100 50 0

0

2

4

6

8 10 12 14 16 18 20 22 24 26

Time/d Time (day)

Fig. 2. Changes of Fe(II) and Fe(III) concentrations during iron reduction in active sludge of different Fe(III): Fe(OH)3 (●); Iron oxide scale (◆); Green ore (△); Red ore (◇); Blank control 1(▼). Error bar indicates standard deviation

1347


2) The initial seed sludge contained considerable amount of ammonia nitrogen (117.9 mg/L) and nitrite (16.2 mg/L), which provide the suitable condition for anaerobic ammonium oxidation (ANAMMOX) bacteria. Ferrous ions contents may promote the performance of ANAMMOX reaction (Zhang et al., 2009). This could be confirmed by the changes of ammonia nitrogen and total nitrogen contents in this study; 3) By inspecting Fig. 2, the total Fe(III) concentration appeared increasing in different degrees after 7 d, especially iron oxide scale and Fe(OH)3 system, showing a sign of a sharp increase in the enrichment culture period. However, with the decrease of ammonia nitrogen and total nitrogen contents, the accumulation of nitrate, which was produced by ANAMMOX process, could promote nitratedependent iron oxidation (Nielsen and Nielsen, 1998; Ratering and Schnell, 2001). This will oxidize the produced ferrous ion to ferric ion. Consequently, ferrous ions were consumed and their concentration was decreased in the system. 3.2. Effects of dissimilatory reduction of different sources on nitrogen and phosphorus removal 3.2.1. Phosphorus removal The concentrations of total phosphate in all the experiments decreased throughout the 25 days of cultivation (Fig. 3). The process with Fe(OH)3 and iron oxide scale showed a high phosphorus removal rate while the phosphate concentration in the blank control remained no change. In particular, in the process with Fe(OH)3 and iron oxide scale, the total phosphate decreases sharply at the initial period of 0 to 2 days, followed by a continuous decrease till the very low concentration of less than 0.5mg/L at the end. Relatively, green ore and red ore systems showed low efficiency of total phosphorus removal and the final concentration of total phosphorus in green ore and red ore system was 2.48 and 4.54 mg/L, respectively. This may be the result to reflect the low iron reducing efficiency in the green ore and red ore systems. In addition, by inspecting Fig.2, it seems clear that the concentration changes of total phosphorus had close relation with Fe(II) and Fe(III) concentrations. Higher iron ions concentration leads to a better phosphorus removal rate. Obviously, the proposed materials have the advantage of low cost in comparison with Fe(OH)3. Nielsen et al. (2002) elucidated the reason that phosphorus removal in iron reducing conditions is the result of biological approach by IRB and nonbiological approach. On the one hand, IRB utilizes Fe(III) as sole elector acceptor. The dissociated Fe(II) is accompanied by precipitation of FeHPO4 and sorption of phosphate onto the surface of iron oxides and hydroxides. The residual Fe(III) may hydrolyze to polynuclear complex, such as Fen(OH)m(3n-m)+(n>1, m=3n), which can promote the

1348

phosphorus removal (Nielsen et al., 2002; Xu et al., 2003). On the other hand, part of phosphorus was adsorbed directly by iron oxide and hydroxide particle surface area (Ivanov et al., 2005). In order to investigate the contribution to phosphorus removal between the role of surface adsorption and chemical sedimentation driven by IRB, further analysis was carried out. Blank control of only Fe(III) sources was used to evaluate the phosphate adsorption onto the surface of iron ore particles in the rotating reactors. The relationship between maximum phosphorus removal rate over 7 days of batch cultivation and the surface adsorption of Fe(III) oxide and hydroxide particle is represented by a columnline plotting, showing in Fig. 4. Chemical precipitation of phosphate by ferrous salts had a P/Fe molar ratio from 0.33 to 1.0 (Takacs et al., 2006), which could hypothetically reflect the proportion of phosphorous removal efficiency by IRB. It is known from Fig. 4 that the phosphorus removal rate depends mainly on the activity of IRB, but the surface adsorption of iron oxide particles remains the secondary important role. Based on the data of maximum ferrous production rate,

with

the

order

of

Fe(OH)3>iron

oxide

scale>green ore>red ore, there is a positive correlation with iron contents. According to P/Fe(II) (mol/mol), the ratio of iron oxide scale is higher than other systems, which is up to 0.17. This is because of the stronger ability of surface adsorption by iron oxide scale. This result is consistent to the data in Table 1. At the end of the experimental period, the phosphate content of iron oxide scale was significantly increased, from the initial 0.5∼0.7% to 2.0∼7.1%. Moreover, high ratio of P/Fe could be explained hypothetically by chelation of phosphate with ferrous ions. Conversely, low ratio of P/Fe in the reactor with green ore and red ore could be explained hypothetically by an excess of ferrous ions for phosphorus removal and the formation of insoluble ferrous iron hydroxide, binding less phosphate than ferrous ions (Ivanov et al., 2009). Hence, based on the experimental data on ferrous production rate and phosphate removal rate of this study, iron ore could replace amorphous ferric hydroxide as a source of ferric for IRB activity. 3.2.2. Nitrogen removal Removal of ammonia nitrogen and total nitrogen by different Fe(III) reduction is shown in Fig. 5(a) (b). The concentration of ammonia nitrogen and total nitrogen remained unchanged in blank control. But the effluent of the experimental systems showed a slightly decreasing trend. According to the changes of effluent concentration, the overall denitrification process can be divided into three stages.

20 15 10 5 0

0

5

10

15 20 25 30 Tim e/d Time (day) Fig. 3. Comparison of phosphorous removal efficiency in addition of different Fe(III) during dissimilary reduction:

4.0

30

3.5

0.20

25

3.0

20

2.5 2.0

15

1.5

10

1.0 5

0.5 0.0

0

0.15 ( mol·mol-1) P/FeP/Fe (mol/mol)

25

Maximum production rate/( mg·L-1·d-1) Maximumferrous ferrous production rate (mg/L.d)

Maximun phorsphate removal rate/( mg·L-1·d-1) Maximumtotal total phosphorus removal rate (mg/L.d)

Total phosphorus concentration (mg/L) phosphorus concentration/(mg/L)


0.10

0.05

0.00

Fe(OH)3 Iron oxide scale Green ore Red ore Iron oxide scale Green ore Red ore Fe(OH)3

Fig. 4. Relationship between phosphorus removal and ferrous production rates: Surface adsorption ( ); Chemical sedimentation driven by IRB( ); P/Fe molar rate (○);

Fe(OH)3 (●); Iron oxide scale (◆); Green ore (△); Red ore (◇);Blank control 1(▼).Error bar indicates standard deviation

Maximum ferrous production rate (◆). Error bar indicates standard deviation.

Table1. Composition of iron oxide scale by IRB before and post-interaction Contents, % Si

Composition Fe Initial state

70.5∼72.0

0.8∼1.1

0.3∼0.8

0.8∼1.8

0.5∼0.7

Else 25.2

After operation

60.9∼69.8

0.6∼1.9

0.9∼2.0

0.8∼5.5

2.0∼7.1

25.0

Total Totalnitrogen (mg/L) nitrogen/(mg/L)

120 100 80 60 40

0

4

8

12 16 20 T im e(day) /d Time

100

24

28

S

P

(b)

160

(a)

Nitrogen Nitrogenremoval accumulation (mg/L) removal accumulation/(mg/L)

Ammoonia nitrogen (mg/L) Ammonia nitrogen/(mg/L)

140

Ca

140 120 100 80 60 40

0

4

8

12

16

T im e/d Time (day)

20

24

28

(c)

80 60 40 20 0

Fe(OH) oxide scale Green ore Red ore Fe(OH)3 3 Iron Iron oxide scale Green ore Red ore

Fig. 5. Comparison of nitrogen removal efficiency in addition different Fe(III) during dissimilary reduction: Fe(OH)3 (●); Iron oxide scale ( ); Green ore ( ); Red ore ( ); Blank control 1(▼); Ammonia nitrogen ( ); Total nitrogen ( ). Error bar indicates standard deviation

1349


At the period of 0 to 10 days, nitrogen removal of various systems was no significant difference and had a large fluctuation. Among these, Fe(OH)3 system showed better performance. The concentrations of ammonia nitrogen and total nitrogen decreased from initial 121.0 and 137.8 mg/L to 88.13 and 106.42 mg/L, respectively. At the period of 10 to 16 days, denitrification efficiency of different Fe(III) sources shows obvious difference, which was positively correlated with initial iron content. The trend is that the higher the iron content, the faster the nitrogen removal rate. The denitrification performance of the iron oxide scale was significantly higher than Fe(OH)3 system at the remaining periods of the experiment. The final removal rate of ammonia nitrogen and total nitrogen reached 64.55% and 62.50%, respectively. Fig. 5(C) shows that the accumulated ammonia nitrogen and total nitrogen removal of iron oxide scale reaches the maximum values of 75.04 and 80.99 mg/L, respectively. Clement et al., (2005) found that nitrification was coupled to iron reduction in the absence of an initial nitrate pool. This may be true that ferric iron has a similar electrochemical property with oxygen. The standard redox potential of Fe3+/Fe2+ is 770 mV whereas that of oxygen is 816 mV. Consequently, ferric iron can generally accept an electron instead of oxygen if the bacteria have proper enzyme systems. For example, Geobacter sulfurreducens and Aeromonas hydrophila can use ferric iron as an electron acceptor under the anoxic condition (Bond and Lovley, 2003; Pham et al., 2003). In addition, more studies have demonstrated a biological reduction of nitrate, and nitrite was found to take place in activated sludge concomitantly with the oxidation of ferrous iron to ferric iron (Nielsen and Nielsen, 1998). Therefore, the reason of denitrification in activated sludge may be the comprehensive action by various microbial redox reactions. The concentrations of nitrate and nitrite in the reaction process and relationship between NH4+ removal amount and NO2 − production will be the subject of later investigation, but is outside the scope of the current study. 4. Possible mechanisms phosphorus removal by reduction

of nitrogen dissimilatory

and iron

The Gibbs free energy of reactions is the index to judge whether the various types of redox reactions can be occurred. Some probable redox reactions coupled to dissimilatory iron reduction and standard Gibbs free energy from the literature are showed in Table 2. It is noteworthy that all these reactions are likely to occur in the current testing systems in this study. Considering the concentration changes of Fe(II), ammonia nitrogen and total nitrogen in the

1350

tested systems, the speculated mechanism of nitrogen and phosphorus removal by dissimilary iron reduction is illustrated in Fig. 6. IRB could drive environmental carbon cycle, but it is also an important participant in circulation driven by other substances (including C, N, P, S). Fig. 6 reflects the coupling mechanism of dissimilatory iron reduction in N and P cycle. Denitrification may occur under the dissimilary iron reducing conditions in the activated sludge, and the redox process can be divided into three subprocesses. Process one could be Feammox. As early as in 1997, Luther et al., (1997) found that NO3- was reduced to N2 by Mn2O in manganese-rich marine sediments under aerobic and anaerobic conditions, respectively, when the classical nitrification/denitrification pathways were shortcircuited by the oxidation of NH3 and organic nitrogen with manganese oxides (MnO2). Clement et al., (2005) demonstrated that Fe(III) hydroxides were proposed to be the electron acceptor for the oxidation of ammonium to NO2- under Fe-reducing conditions in soil incubation experiments. Later studies observed the production of NO2under iron-reducing conditions by an anaerobic culture from wetland soils (Shrestha et al., 2009; Xu et al., 2011; Yang et al., 2012) and piggery wastewater (Park et al., 2009), in the absence of an initial nitrate pool. Most recently, Yang et al., (2012) verified NH4+ loss coupled to iron reduction under anaerobic conditions in soils through the way of oxidizing NH4+ to N2 by isotopic tracing and calculation of the thermodynamic favorability. Process two is anaerobic ammonium oxidation. The ANAMMOX has been widely studied and it requires NO2- for the anaerobic oxidation of NH4+. The mechanism had been confirmed. But problems still exist in starting time of the ANAMMOX reactor, influencing factors and microbial properties etc. Process three is nitrate-dependent iron oxidation. A biological reduction of nitrate was found to be used as electron acceptor, rather than oxygen, for the oxidation of the ferrous iron to ferric iron. Studies have reported in literature (Nielsen and Nielsen, 1998) that this process was present in different types of activated sludge treatment plants. The main product of Fe(II)-dependent nitrate removal was most probably dinitrogen, as no accumulation of ammonia, nitrous oxide, or nitrite could be observed (Nielsen and Nielsen, 1998). Moreover, a number of microorganisms have been described, which can couple iron oxidation to nitrate reduction in the absence of oxygen and light (Lack et al., 2002; Muehe et al., 2009). A recent study has focused on the mechanisms of iron oxidation by anaerobic nitrate reducing bacteria. For example, Carlson et al., (2012) proposed a mechanistic model by balancing electron uptake and detoxification.


Table 2. ΔGθ of probable redox reaction Reaction Type

Reaction +

Dissimilatory iron reduction ANAMMOX Feammox Nitrate-dependent Iron oxidation

-

2+

FeOOH+3H +e →Fe +H2O NH4++NO2-→N2+2H2O NH4++6Fe3++10H+→NO2-+6Fe2++2H2O 10Fe2++2NO3+24H2O→10Fe(OH)3+N2+18H+

N org

Fixation

Ammonifications

ΔGθ/kJ·m ol-1 -64.6

Possible

Clement et al., 2005

-335 -24.3 -75.9

Possible Possible Possible

Broda, 1997 Zhang et al., 2009 Nielsen and Nielsen, 1998

(3n-m)+

Fen(OH)m Hydrolysis

I

Denitrification

NO 3-

NO 2

or

-

II

III N2 Fe(II)

References

FeOOH

NH 4+

N2

Fe(III)

Possibility

Fe(II) (dissoloved) HPO 42-

IV Fe(III) (dissoloved) III

IV

FeHPO4

Fig. 6. Proposed mechanism of multi-component interactions by microorganisms I: Feammox; II: Anammox; III: Nitrate-dependent iron(II) oxidizers; IV: non-biological ways

It should be noted that the activated sludge system of municipal WWTPs is a complicated food chain consisted of a huge variety of microorganisms and the nutrients in the wastewater. For better simulation and analysis of the denitrification process in this system as well as the nitrogen cycle paths, further study should be conducted based on the enrichment culture of different genus, tracking changes in the characteristic index of concentration, such as isotopic tracer etc. 5. Conclusions The utilization efficiency of different Fe(III) sources shows significant difference in microbial dissimilatory iron reduction. Although the amorphous and dissolved iron (such as Fe(OH)3) of large specific surface area has advantage in utilization, the iron oxide scale is more easily reduced by specific microorganism. Thus it can be used to replace the traditional Fe(OH)3 as Fe(III) source and in the same time to achieve the aim of using “waste” for “waste control”. The phosphorus removal efficiency is positively related to the produced ferrous concentration, which is determined by the iron content of the material, with the order of Fe(OH)3>iron oxide scale>green ore>red ore. After 25 days incubation under strict anaerobic conditions, TP removal rate reached 98.0%, 99.1%, 91.2% and 73.4% in the case of Fe(OH)3, iron oxide scale, green ore and red ore being used, respectively.

According to inspecting the standard Gibbs free energy change of each redox reaction, the anaerobic dissimilatory iron reduction may be coupled to ammonia oxidation. This leads to integrated biochemical reactions, which include ANAMMOX and nitrate dependent Fe(II) oxidation. Nitrogen removal in the system may be the results of complex interaction. Iron oxide scale showed better performance with ammonia nitrogen and total nitrogen removal rate of 64.55% and 62.50%, respectively. However, NH4+-N and TN removal exhibits a removal hysteresis compared with total phosphorus. Acknowledgements The authors wish to thank the financial support received from the National Nature Foundation of China (51068015) and the major project of the National Water Pollution Control and Management Technology (2012ZX07201-00505-02).

References Benner S.G., Hansel C.M., Wielinga B.W., Baarber T.M., Fendorf S., (2002), Reductive dissolution and biomineralization of iron hydroxide under dynamic flow conditions, Environmental Science and Technology, 36, 1705-1711. Bond D.R., Lovley D.R., (2003), Electricity production by geobacter sulfurreducens attached to electrodes, Applied Environmental Microbiology, 69, 1548-1555. Botton S,, van Harmelen M., Braster M., Parsons J.R., Roling W.F.M., (2007), Dominance of Geobacteraceae in BTX- degrading enrichments from

1351


an iron-reducing aquifer, FEMS Microbiology Ecology, 62, 118–130. Broda E., (1977), Two kinds of lithotophs missing in nature, Journal of Basic Microbiology, 17, 491-493. Carlson H.K., Clark I.C., Melnyk R.A., Coates J.D., (2012), Toward a mechanistic understanding of anaerobic nitrate-dependent iron oxidation: balancing electron uptake and detoxification, Microbiology, 3, 16, doi: 10.3389/fmicb.2012.00057. Clement J.C., Junu 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 Biochemistry, 37, 23232328. Ivanov V., Kuang S., Stabnikov V., Guo C., (2009), The removal of phosphorus from reject water in a municipal wastewater treatment plant using iron ore, Chemical Technology and Biotechnology, 84, 78-82. Ivanov V., Lim J.J.W., Stabnikova O., Gin, K.Y.H., (2010), Biodegradation of estrogens by facultative Process anaerobic iron-reducing bacteria, Biochemistry, 45, 284–287. Ivanov V., Stabnikov V., Zhuang W.Q., Tay J.H., Tay S.T.L., (2005), Phosphate removal from the returned liquor of municipal wastewater treatment plant using Journal of Applied iron-reducing bacteria, Microbiology, 98, 1152–1161. Lack J.G., Chaudhuri S.K., Chakraborty, R., Achenbach L.A., Coates J.D., (2002), Anaerobic biooxidation of Fe(II) by dechlorosoma suillum, Microbiology Ecology, 4, 424–431. Li F.B., Li X.M., Zhou S.G., Zhuang L., Cao F., Huang D.Y., Xu W., Liu T.X., Feng C.H., (2010), Enhanced reductive dechlorination of DDT in an anaerobic system of dissimilatory iron-reducing bacteria and iron oxide, Environmental Pollution, 158, 1733-1740. Li J., Zhang Y., Shi Z., (2009), Microbial immobilization filler with nano-attapulgite composite hydrophilicity polyurethane foam, China Patent, No. zl2009 1 0117393.3[Patent].2009-07-22. Lovley, D.R., Phillips, E.J.P., (1986), Organic matter mineralization with the reduction of ferric iron in Applied Environmental anaerobic sediments, Microbiology, 51, 683-689. Lovely D.R., Phillips E.J.P., (1988), Novel mode of microbial energy metabolism: organic carbon oxidation coupled to dissimilatory reduction of iron or manganese, Applied Environmental Microbiology, 54, 1472-1480. Luan F., Xie L., Li J., Zhou Q., (2009), Reduction of nitrobenzene by iron oxides bound Fe(II) system at different pH values, Environmental Science, 30, 19371941. Luther G.W., Sundby B., Lewis B.L., Brendel P.J., Silverberg N., (2007), Interactions of manganese with the nitrogen cycle: alternative pathways to dinitrogen, Geochimca et Cosmochim Acta, 61, 4043–4052. Muehe E.M., Gerhardt S., Schink B., Kappler A., (2009), Ecophysiology and the energetic benefit of mixotrophic Fe(II) oxidation by various strains of nitrate-reducing bacteria, FEMS Microbiology Ecology, 70, 335–343. Nielsen JL., Juretschko S., Wagner M., Nielsen PH., (2002), Abundance and phylogenetic affiliation of iron

1352

reducers in activated sludge as assessed by fluorescence in situ hybridization and microautoradiography, Applied Environmental Microbiology, 68, 4629-4636. Nielsen J.L., Nielsen P.H., (1998), Microbial nitratedependent oxidation of ferrous iron in activated sludge, Environmental Science and Technology, 32, 3556-3561. Orbeci C., Untea I., Stanescu R., Segneanu A.E., Craciun M.E., (2012), a new photo-Fenton procedure applied in oxidative degradation of organic compounds from Environmental Engineering and wastewater, Management Journal, 11, 141-146. Park W., Nam Y.K., Lee M.J., Kim T.H., (2009), Anaerobic ammonia-oxidation coupled with Fe3+ reduction by an anaerobic culture from a piggery wastewater acclimated to NH4+/Fe3+ medium, Biotechnology and Bioprocess Engineering, 14, 680685. Pham C.A., Jung S.J., Phung N.T., Lee, J., Chang I.S., Kim B.H., Yi H., Chun J., (2003), A novel electrochemically active and Fe(III)-reducing bacterium phylogenetically related to aeromonas hydrophila, isolated from a microbial fuel cell, FEMS Microbiology Letters, 223, 129-134. Ratering S., Schnell S., (2001), Nitrate-dependent iron(II) oxidation in paddy soil, Environmental Microbiology, 3, 100–109. Shrestha J., Rich J.J., Ehrenfeld J.G., (2009), Oxidation of ammonium to nitrite under iron-reducing conditions in wetland soils: laboratory, field demonstration and push-pull rate determination, Soil Science, 174, 156164. Stabnikov V.P., Tay S.T.-L., Tay D.-Kh., Ivanov VN., (2004), Effect of iron hydroxide on phosphate removal during anaerobic digestion of activated sludge, Applied Biochemistry and Microbiology, 40, 376–380. Takacs I., Murthy S., Smith S., McGrath M., (2006), Chemical phosphorus removal to extremely low levels: experience of two plants in the Washington, DC area, Water Science and Technology, 53, 21–28. WWDA, (2002), Water and Wastewater Detection Fourth Edition, China Analysis Methods, Environmental Science Press. Xu F., Luo J., Ling D., (2003), Present and prospects of the removal of phosphorus from wastewater chemically, Industrial Water Treatment, 23, 18-20. Xu D.F., Li Y.X., Xu X.H., Zhao X.L., Fang H., (2011), Effect of microbial activity in the rhizosphere of wetland plants on removal of total organic carbon and nitrogen from wastewater, Environmental Engineering and Management Journal, 10, 781-786. 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, 538541. Zhang L., Zheng P., Hu A., (2009), Effect of ferrous ion on the performance of an anammox reactor, Acta Scientiae Circumstantiae, 29, 1629-1634. Zhu W., Zang H., Wu F., (2011), The kinetic characteristics of the microbial reduction of goethite, China Environmental Science, 31, 820-827.


July 2013, Vol.12, No. 7, 1353-1358



WHOLE CELL BIOREPORTER FOR THE ESTIMATION OF OIL CONTAMINATION Chuan Li1,2, Dayi Zhang2, Yizhi Song1, Bo Jiang1, Guanghe Li1, Wei E. Huang2 1

School of Environment, Tsinghua University, Beijing 100084, P.R. China Kroto Research Institute, University of Sheffield, Broad Lane, Sheffield, S3 7HQ, UK

2

Abstract Oil contamination has been one of the most common environmental pollution due to its threat to both human health and ecosystems. It is difficult for the application of conventional chemical analysis to detect oil contamination in-situ, estimate bioavailability and rapidly assess the environmental risk. In this study, two bioreporters were employed for the rapid and direct detection of oil content and genotoxicity in a cutting oil contaminated site in China. Acinetobacter baylyi ADPWH_alk and ADPWH_recA are chromosomally-based bioreporter for the detection of alkanes and genotoxicity separately. The reporter gene cluster luxCDABE enabled cells to express bioluminescence in the presence of contaminants. Calibration curve of bioreporter response to oil in the soils illustrated a log-log linear relationship between oil content and bioluminescence expression, indicating that whole cell bioreporter can be used as a semi-quantitative analysis method to estimate the oil contamination. The oil contamination in soils detected by ADPWH_alk (5760 and 1090 mg/kg) were comparable with those obtained by GC/MS (7070 and 1490 mg/kg). Furthermore, the genotoxicity bioreporter ADPWH_recA showed that the ecological genotoxicity of the soil samples were equivalent to 33-36 mg mitomycin C per kilogram soils. The results suggested that bioreporter sensing could be a rapid, simple, low-costly and in-situ detection method with unique ability for the detection of toxicity and bioavailability, and it can be a useful and complementary tool to chemical analysis. Key words: Acinetobacter baylyi ADP1, alkane, bioreporter, cutting oil, genotoxicity, mineral oil, oil contamination Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction As an important and essential energy source, petroleum plays important roles in modern industry. However, during the process of oil production, transportation, storage and disposal, significant oil contamination events pose environmental risk. Since the first oil spill occurred in 1969, there have been more than 40 large-scale oil spills, including the United States Alaska Exxon Valdez Oil Spill in 1989 (Bence et al., 1996; Bragg et al., 1994) and the Mexico Gulf Deepwater Horizon Oil Spill in 2010 (Camilli et al., 2010). With rapid development of Chinese industry and municipal construction, oil spill and oil contaminated sites are increasingly serious, 

causing a major threat to people’s health and ecological system. Chinese oil contamination is so serious that one large-scale oil spill occurred every four days since 2006 according to the statistics of the State Oceanic Administration (Zhang et al., 2013). Furthermore, oil contaminated sites in oil field or oilprocessing factories also result in large areas of soil pollution, and have drawn public attentions (Peterson et al., 2003; Piatt et al., 1990). Many bacteria have been found to degrade nalkane in oil contaminated sites, including Acinetobacter, Alcaligenes (Lal and Khanna, 1996), Alcanivorax (Hara et al., 2003), Arthrobacter (Radwan et al., 1996), Geobacillus (Feng et al., 2007), Bacillus (Chaerun et al., 2004; Kato et al.,

Author to whom all correspondence should be addressed: E-mail: : [email protected]; Phone: +44 (0)114 2225796; Fax: +44 (0)114 2225701

Li et al./Environmental Engineering and Management Journal 12 (2013), 7, 1353-1358

2001), Brachybacterium (Yan, 2006), Burkholderia (Yuste et al., 2000), Desulfatibacillum (CravoLaureau et al., 2004; Cravo-Laureau et al., 2005), Dietzia (von der Weid et al., 2007; Yumoto et al., 2002), Geobacillus (Wang et al., 2006), Gordonia (Kotani et al., 2003), Marinobacter (Bonin et al., 2004; Doumenq et al., 2001), Mycobacterium (Churchill et al., 1999; van Beilen et al., 2002), Paracoccus (Chaerun et al., 2004; Zhang et al., 2004), Planococcus (Engelhardt et al., 2001), Pseudomonas (Koch et al., 1991; Naik and Sakthivel, 2006), Rhodococcus (Andreoni et al., 2000; Kunihiro et al., 2005; van Beilen et al., 2002) and Thermooleophilum (Zarilla and Perry, 1984). Though gene regulation systems in these strains are potentially used to construct various bioreporters to sense n-alkane, there are only a few alkane bioreporters have been developed (Alkasrawi et al., 1999; Minak-Bernero et al., 2004; Sticher et al., 1997). However, most of those bioreporters can only sense short and medium (C5-C12) carbon chain of alkanes and alkenes. Amongst those alkane degraders, Acinetobacter baylyi ADP1 can utilise alkanes with carbon lengths ranging from 12 up to 36 and the gene regulation for alkane degradation is well characterised (Ratajczak et al., 1998a; Ratajczak et al., 1998b; Throne-Holst et al., 2007). Acinetobacter alkane bioreporter ADPWH_alk was constructed previously and was able to sense a broad range of carbon chain length alkanes from C7 to C36 (Zhang et al., 2012). ADPWH_alk was induced by alkanes in short time (about 30 min) and the response is semiquantitative. In contrast, other previous reported bioreporters could only be induced after 10 hours (Ratajczak et al., 1998a). In addition, ADPWH_alk was also functional in fresh water, seawater (Zhang et al., 2013) and soil samples. In this study, Acinetobacter baylyi ADPWH_alk bioreporter was used to demonstrate its practical application to estimate the oil contaminated soil samples from a cutting oil contaminated site. 2. Material and methods 2.1. Site description Four oil contaminated soil samples were taken from a metal processing factory that utilizes mineral oil as lubricating oil in Zhejiang Province, China, which was established as an automobile part manufacturing facility since 1996. Two soil samples, SB-1-S and SB-2-S, were taken from surface layers of the contaminated site and other two samples, HA-3-S and HA-3-S, were from a nearby uncontaminated site. Site investigation indicated that petroleum and cutting oil contamination had occurred since the factory was setup. Across 200 m2 contaminated area, the main pollutants were comprised of mineral oil and surfactant.

1354

The average concentration of the total petroleum pollutants was higher than 5000 mg/kg soil and the maximum concentration was about 25,000 mg/kg soil. This severe contaminated soil was located just above groundwater water table 0.3 m to 1.0 m, leading to the concentration of petroleum and cutting oil plume in groundwater greater than 0.6 mg/L covering 500 m2. For bioreporter detection, multiple samples ranging from 5 to 200 mg were randomly taken from each soil samples and four replicates of each sample were carried out. 2.2. Cell culture and pre-treatment ADPWH_alk, a chromosomally-based Acinetobacter bioreporter (Zhang et al., 2012), was used to estimate the crude oil content of samples. Another bioreporter, ADPWH_recA (Song et al., 2009), was used to evaluate genotoxicity and convert to standard genotoxic chemicals, mitomycin C (MMC). After growth in lysogeny broth (LB) medium at 30ºC overnight, the cells of both bioreporters were harvested by centrifugation at 3000 rpm for 10 minutes at 4ºC, and subsequently washed by 0.85% NaCl solution and re-suspended in the same volume of saline solution. The bioreporters were stored at 4ºC and ready for use. The bioreporters can retain their sensing activities for at least 8 weeks (Song et al., 2009). 2.3. Bioreporter detection All the bioreporter detection was carried out in LB medium. Briefly, 1 mL ADPWH_alk or ADPWH_recA stock solution (about 108 CFU/mL) was centrifuged at 4000 rpm for 5 minutes and resuspended in 10 mL LB medium. Daqing crude oil, which is used as a standard crude oil in China, was added into 10 mL volumes of chloroform to give the following oil content: 0, 0.1, 1, 5, 10, 25, 50 and 100 mg. Then the oil-chloroform mixtures were added into 20 g standard soils, which were volatilized at 30ºC for 24 hr to remove chloroform residue from the soil, to give a final crude oil content of 0%, 0.5%, 1%, 2.5%, 5%, 10%, and 15% (weight oil /weight soil). Four hundred milligram of each oil-soil sample were mixed with 5 mL deionized water and exposed to 40 kHz ultrasound for 300 seconds. After settlement for 10 minutes, 900 μL of the soil/water mixture supernatant was mixed with 100 μL ADPWH_alk or ADPWH_recA cell suspensions, before the samples were loaded onto the microplate reader. Pure alkanes (Sigma-Aldrich, Co.) with different carbon chain length: hexane (13.1 µL), heptane (14.7 µL), octane (16.2 µL), dodecane (22.7 µL), tetradecane (26.0 µL), octadecane, (25.5 mg), tetracosane (33.9 mg), triacontane (42.3 mg), hexatriacontane (50.7 mg) and tetracontane (56.3

Whole cell bioreporter for the estimation of oil contamination

mg), were dissolved in 100 ml deionized water respectively. After homogenization using a 40K-Hz ultrasound for 30 seconds, the n-alkanes emulsion stock solutions were obtained with final concentration of 1.0 mM. The emulsion treatment enabled alkanes to homologically distribute in water although they were insoluble. 2.4. Bioluminescence measurement The 200 μL mixtures of ADPWH_alk/ADPWH_recA from above were added into each well of a black clear-bottom 96-well microplate (Corning, USA). Three measurement replicates were carried out for each sample. The microplate was incubated at 30°C and monitored for 4 hr. The bioluminescence was measured every 10 minutes using a SpectraMax M5 multimode microplate reader (Molecular Devices, USA), equipped with SoftMax Pro analysis software. Before each measurement, 5 minutes of vertical shaking was used for oil inducer dispersion and cell growth. 2.5. Data analysis Induced bioluminescence of ADPWH_alk and ADPWH_recA was obtained by averaging five bioluminescence measurements between 180 minutes and 210 minutes. The bioluminescence response ratio was calculated by dividing induced bioluminescence by the original bioluminescence (time = 0). The relative bioluminescence response ratio was evaluated by dividing the induced bioluminescence by the bioluminescence of negative control (non-induced) samples. OD600 was taken twice before and after bioluminescence detection, to determine cell growth conditions.

3. Results 3.1. Detection of a broad range of carbon chain alkenes and alkanes (C7-C36) It has been shown that ADPWH_alk responded to n-alkane with carbon chain length from 7 to 36 in minimal medium with 20 mM sodium succinate as the sole carbon source (Zhang et al., 2012). In this study, LB medium was employed for the bioreporter detection of both oil and genotoxicity. In consistent with previous report, the results showed that ADPWH_alk was able to detect alkanes with carbon chain length ranging from C7 to C36 in LB medium as well (Fig. 1). Though Acinetobacter bayiyi ADP1 can only utilise alkanes with carbon chain greater than 12 as sole carbon source (Ratajczak et al., 1998a; Wentzel et al., 2007), ADPWH_alk can sense a broader range of alkenes from C7 to C36. ADPWH_alk induction ratio varied in response to alkanes with different lengths of carbon chain (Fig. 1), which is in a good agreement with the previous observation when ADPWH_alk was induced in minimal medium (Zhang et al., 2012). The results indicated that the induction ratios of alkanes with C12, C18 and C24 were significantly higher than those with other carbon chain lengths. It suggested that alkanes with specific carbon chain lengths could affect the complex of AlkR-promoter-RNAP and promote alkM transcription process. 3.2. Calibration of ADPWH_alk responding to crude oil in soil As demonstrated previously, the increase of the induction ratio of ADPWH_alk to Daqing crude oil is proportional to the oil content in water (Zhang et al., 2012).

2.6. Chemical analysis The chemical analysis of crude oil was carried out by gravimetric and GC/MS methods, as previously described (Zhang et al., 2012). Briefly, 200 mL of marine water and 10.0 g sediment samples were extracted by 50 mL chloroform solvent. The extracts were passed through an anhydrous sodium sulphate column to remove water and suspended solids. The total hydrocarbon content was determined gravimetrically after the solvent was evaporated. Subsequently, 30-80 mg of the crude oil/soil extract was dissolved in 5 mL of n-hexane and loaded onto a glass column. The saturated fraction was eluted using 30 mL of 95% n-hexane, concentrated to 1 mL, and analysed using an Agilent 7890A GC coupled with an Agilent 5975C MS operated in the electron impact mode. The GC/MS was calibrated using solutions of n-C9 to n-C40 alkane standard compounds at seven different concentrations and 5α-Androstane was used as the internal standard.

Fig. 1. ADPWH_alk response to alkanes with different carbon chain lengths from 6 to 36 in LB medium. The alkanes were emulsified by ultrasound and the final concentration of each alkane or alkene was 100 µM. pvalue (against control) < 0.001 for the t-test of induction ratios

1355


A similar trend was also observed when ADPWH_alk was applied to sense the standard crude oil in soil samples in the range of 1-15% (weight oil /weight soil) (Fig. 2).

1). In the genotoxicity estimation, mitomycin C (MMC) was used as a standard genotoxin and the genotoxicity effect was defined as mg MMC per kilogram soils. It indicated that the soil samples containing cutting oils exhibited 10 times more genotoxic effect than the control soil samples (Table 1). 4. Discussion

Fig. 2. Calibration curve for ADPWH_alk response to crude oil in soils

At the time (200 to 240 min) of the maximal induction, a calibration relationship between the bioluminescence expression of ADPWH_alk and the crude oil concentration in soils was established in Fig. 2. Although ADPWH_alk sensed individual alkane differently (Fig. 1) and the crude oil was consist of alkane mixtures with different compositions, the induction peak times were the same and an overall induction pattern in Fig. 2 indicated that there was a clear relationship between the bioreporter’s bioluminescence expression and the content of crude oil in soils. 3.3. Detection of cutting oils in soil samples The oil content of the cutting oil contaminated soils was measured by both bioreporter ADPWH_alk and chemical GC/MS. With the calibration curve of ADPWH_alk induction ratio against crude oil content in soils (Fig. 2), the cutting oils content in four soil samples can be estimated. At the same time, the oil content of the cutting oils was also measured using GC/MS. The results from bioreporter detection and GC/MS indicated comparable data for the cutting oil content in soil samples. Table 1 shows that the oil content estimated by bioreporter ADPWH_alk (5760 and 1090 mg/kg) was slightly lower than the results obtained by GC/MS (7070 and 1490 mg/kg), which may be due to the fact that bioreporter actually detected the bioavailable portion of oil in the soil samples whereas GC/MS extracted and measured the total amount of oils including inert parts binding to the soil or clay particles. The genotoxicity bioreporter ADPWH_recA (Song et al., 2009) was also applied to the soil samples and the cutting oil contents found in soil and genotoxicity are in good agreement (Table

1356

GC/MS is an accurate analytical technique, whereas the bioreporter approach could be complementary to GC/MS for rapid sample screening. The chemical analysis of all soil samples by GC/MS (USEPA, 1996; USEPA, 2006) methods followed USEPA 8260C and 8015D procedures. The comparable data of chemical and bioreporter detection indicated that bioreporter can be a rapid and semi-quantitative tool for measurement of oil contaminated soil samples, complementary to more accurate and reliable chemical analysis. Bioreporters employing live organisms as sensing agents could be complementary tools to chemical analysis while representing a few advantages: rapid, simple, cheap, in-situ and importantly providing information of bioavailability and toxicity. Bioavailability and toxicity are main concerns for environmental risk assessment and management which directly link environmental contamination to human health risk. In natural environment, the majority samples are mixtures of complex contaminants. The additive, antagonistic, and synergistic effects caused by complex physical or chemical interactions would make the risk assessment unpredictable if chemical analysis (e.g. GC-MS or HPLC) alone is applied. Chemical analysis can accurately measure individual compounds but it is difficult to distinguish inert or active portion of a compound and effect of the mixed compounds. In contrast, bioreporters targets active or bioavailable portions of the contaminants and it directly links contamination to biological effects and make the assessment relevant to health risk. It has been a challenge to construct bioreporters with high reliability, robustness and reproducibility. With recent advances in molecular cell biology, synthetic biology and nanotechnology, bioreporters could be conferred with new functions which enhanced the sensing performances (Zhang et al., 2011), which would reduce the gap between laboratory test and real-world applications. Compared with standard USEPA methods, the bioreporter developed in this study for the detection of crude oil contamination in marine water and sediment samples has many distinguishing features and possibilities for future application. Firstly, the bioreporter method needs only 1 mL water sample and 1.0 g sediment sample for measurement, whilst the recommended USEPA GC/MS methods require at least 500 mL (Wang and Fingas, 1997).

Whole cell bioreporter for the estimation of oil contamination

Table 1. Oil content and genotoxicity detection of cutting oil contaminated soils Soil Sample

SB-1-S SB-2-S HA-3-S HA-3-S

GC-MS Method Oil Content (mg/kg) 7070 1490 NDa NDa

Bioreporter Method Oil content a Genotoxicityb (mg/kg) 5760 33 1090 36 45 3 47 3

a

Estimated values according to the oil-content/soil calibration curve in Fig. 2; bGenotoxicity means equivalent to standard genetic toxic compound mitomycin C (MMC).

Secondly, qualitative results from the bioreporter measurement are obtained in less than 0.5-4 hours, whilst 48 hours is usually needed for the GC/MS method. Thirdly, bioreporter method can be applied to in situ detection and online monitoring with minimal sample pre-processing, whilst chemical extraction is usually required before GC/MS. Finally, bioreporter method is low-costly, simple and easy to operate 5. Conclusions In summary, this study demonstrated that bioreporter detection could be a rapid, cheap and semi-quantitative method to estimate oil contamination in water, seawater and soils, and meet the requirement of in-situ monitoring in the field. Compared with chemical analysis, bioreporter method has unique advantages in terms of detection for bioavailability and toxicity which provide vital information for environmental management and assessment of the impact to human health and ecosystems. Acknowledgements The work is supported by EU grant ECOWATER (RFCRCT-2010-00010). The authors would also like to thank the Independent Scientific Research Plan of Tsinghua University (Grant No. 2009THZ0), China MOST (Grant No. 20121887929), and the Doctoral Short-Term VisitingAbroad Foundation of Tsinghua University, Beijing, P.R. China.

References Alkasrawi M., Nandakumar R., Margesin R., Schinner F., Mattiasson B., (1999), A microbial biosensor based on Yarrowia lipolytica for the off-line determination of middle-chain alkanes, Biosensors & Bioelectronics, 14, 723-727. Andreoni V., Bernasconi S., Colombo M., van Beilen J.B., Cavalca L., (2000), Detection of genes for alkane and naphthalene catabolism in Rhodococcus sp strain 1BN, Environmental Microbiology, 2, 572-577. Bence A.E., Kvenvolden K.A., Kennicutt M.C., (1996), Organic geochemistry applied to environmental assessments of Prince William Sound, Alaska, after the Exxon Valdez oil spill - A review, Organic Geochemistry, 24, 7-42.

Bonin P., Cravo-Laureau C., Michotey V., Hirschler-Rea A., (2004), The anaerobic hydrocarbon biodegrading bacteria: An overview, Ophelia, 58, 243-254. Bragg J.R., Prince R.C., Harner E.J., Atlas R.M., (1994), Effectiveness of bioremediation fro the Exxon-Valdez oil-spill, Nature, 368, 413-418. Camilli R., Reddy C.M., Yoerger D.R., Van Mooy B.A.S., Jakuba M.V., Kinsey J.C., McIntyre C.P., Sylva S.P., Maloney J.V., (2010), Tracking hydrocarbon plume transport and biodegradation at deepwater horizon, Science, 330, 201-204. Chaerun S.K., Tazaki K., Asada R., Kogure K., (2004), Bioremediation of coastal areas 5 years after the Nakhodka oil spill in the Sea of Japan: isolation and characterization of hydrocarbon-degrading bacteria, Environment International, 30, 911-922. Churchill S.A., Harper J.P., Churchill P.F., (1999), Isolation and characterization of a Mycobacterium species capable of degrading three- and four-ring aromatic and aliphatic hydrocarbons, Applied and Environmental Microbiology, 65, 549-552. Cravo-Laureau C., Matheron R., Cayol J.L., Joulian C., Hirschler-Rea A., (2004), Desulfatibacillum aliphaticivorans gen. nov., sp nov., an n-alkane- and n-alkene-degrading, sulfate-reducing bacterium, International Journal of Systematic and Evolutionary Microbiology, 54, 77-83. Cravo-Laureau C., Grossi V., Raphel D., Matheron R., Hirschler-Rea A., (2005), Anaerobic n-alkane metabolism by a sulfate-reducing bacterium, Desulfatibacillum aliphaticivorans strain CV2803, Applied and Environmental Microbiology, 71, 34583467. Doumenq P., Aries E., Asia L., Acquaviva M., Artaud J., Gilewicz M., Mille G., Bertrand J.C., (2001), Influence of n-alkanes and petroleum on fatty acid composition of a hydrocarbonoclastic bacterium: Marinobacter hydrocarbonoclasticus strain 617, Chemosphere, 44, 519-528. Engelhardt M.A., Daly K., Swannell R.P.J., Head I.M., (2001), Isolation and characterization of a novel hydrocarbon-degrading, Gram-positive bacterium, isolated from intertidal beach sediment, and description of Planococcus alkanoclasticus sp nov, Journal of Applied Microbiology, 90, 237-247. Feng L., Wang W., Cheng J., Ren Y., Zhao G., Gao C., Tang Y., Liu X., Han W., Peng X., Liu R., Wang L., (2007), Genome and proteome of long-chain alkane degrading Geobacillus thermodenitrificans NG80-2 isolated from a deep-subsurface oil reservoir, Proceedings of the National Academy of Sciences of the United States of America, 104, 5602-5607. Hara A., Syutsubo K., Harayama S., (2003), Alcanivorax which prevails in oil-contaminated seawater exhibits broad substrate specificity for alkane degradation, Environmental Microbiology, 5, 746-753. Kato T., Haruki M., Imanaka T., Morikawa M., Kanaya S., (2001), Isolation and characterization of long-chainalkane degrading Bacillus thermoleovorans from deep subterranean petroleum reservoirs, Journal of Bioscience and Bioengineering, 91, 64-70. Koch A.K., Kappeli O., Fiechter A., Reiser J., (1991), Hydrocarbon assimilatin and biofurfactant production in Psuedomonas-Aeruginosa mutants, Journal of Bacteriology, 173, 4212-4219. Kotani T., Yamamoto T., Yurimoto H., Sakai Y., Kato N., (2003), Propane monooxygenase and NAD(+)dependent secondary alcohol dehydrogenase in

1357


propane metabolism by Gordonia sp strain TY-5, Journal of Bacteriology, 185, 7120-7128. Kunihiro N., Haruki M., Takano K., Morikawa M., Kanaya S., (2005), Isolation and characterization of Rhodococcus sp strains TNP2 and T12 that degrade 2,6,10,14-tetramethylpentadecane (pristane) at moderately low temperatures, Journal of Biotechnology, 115, 129-136. Lal B., Khanna S., (1996), Degradation of crude oil by Acinetobacter calcoaceticus and Alcaligenes odorans, Journal of Applied Bacteriology, 81, 355-362. Minak-Bernero V., Bare R.E., Haith C.E., Grossman M.J., (2004), Detection of alkanes, alcohols, and aldehydes using bioluminescence, Biotechnology and Bioengineering, 87, 170-177. Naik P.R., Sakthivel N., (2006), Functional characterization of a novel hydrocarbonoclastic Pseudomonas sp strain PUP6 with plant-growthprornoting traits and antifungal potential, Research in Microbiology, 157, 538-546. Peterson C.H., Rice S.D., Short J.W., Esler D., Bodkin J.L., Ballachey B.E., Irons D.B., (2003), Long-term ecosystem response to the Exxon Valdez oil spill, Science, 302, 2082-2086. Piatt J.F., Lensink C.J., Butler W., Kendziorek M., Nysewander D.R., (1990), Immediate impact of the Exxon Valdez oil-spill on marine birds, Auk, 107, 387397. Radwan S.S., Sorkhoh N.A., Felzmann H., ElDesouky A.F., (1996), Uptake and utilization of n-octacosane and n-nonacosane by Arthrobacter nicotianae KCC B35, Journal of Applied Bacteriology, 80, 370-374. Ratajczak A., Geissdorfer W., Hillen W., (1998a), Expression of alkane hydroxylase from Acinetobacter sp. strain ADP1 is induced by a broad range of nalkanes and requires the transcriptional activator AlkR, Journal of Bacteriology, 180, 5822-5827. Ratajczak A., Geissdorfer W., Hillen W., (1998b), Alkane hydroxylase from Acinetobacter sp. strain ADP1 is encoded by alkM and belongs to a new family of bacterial integral-membrane hydrocarbon hydroxylases, Applied and Environmental Microbiology, 64, 1175-1179. Song Y.Z., Li G.H., Thornton S.F., Thompson I.P., Banwart S.A., Lerner D.N., Huang W.E., (2009), Optimization of bacterial whole cell bioreporters for toxicity assay of environmental samples, Environmental Science & Technology, 43, 7931-7938. Sticher P., Jaspers M.C.M., Stemmler K., Harms H., Zehnder A.J.B., vanderMeer J.R., (1997), Development and characterization of a whole-cell bioluminescent sensor for bioavailable middle-chain alkanes in contaminated groundwater samples, Applied and Environmental Microbiology, 63, 40534060. Throne-Holst M., Wentzel A., Ellingsen T.E., Kotlar H.K., Zotchev S.B., (2007), Identification of novel genes involved in long-chain n-alkane degradation by Acinetobacter sp strain DSM 17874, Applied and Environmental Microbiology, 73, 3327-3332. USEPA, (1996), US EPA Method 8015 GRO (Solid Waste/Ground Water Gasoline Range Organics by GC-FID) using Direct Injection on SLB-5ms, On line at: http://www.sigmaaldrich.com/content/dam/sigmaaldrich/docs/Supelco/Application_Notes/1/10994_US %20EPA%20Method%208015%20GRO%20%20SLB-5ms.pdf

1358

USEPA, (2006), US EPA method 8260C, Volatile Organic Compounds By Gas Chromatography/Mass Spectrometry (GC/MS), On line at: http://www.epa.gov. van Beilen J.B., Smits T.H.M., Whyte L.G., Schorcht S., Rothlisberger M., Plaggemeier T., Engesser K.H., Witholt B., (2002), Alkane hydroxylase homologues in Gram-positive strains, Environmental Microbiology, 4, 676-682. von der Weid I., Marques J.M., Cunha C.D., Lippi R.K., dos Santos S.C.C., Rosado A.S., Lins U., Seldin L., (2007), Identification and biodegradation potential of a novel strain of Dietzia cinnamea isolated from a petroleum-contaminated tropical soil, Systematic and Applied Microbiology, 30, 331-339. Wang L., Tang Y., Wang S., Liu R.-L., Liu M.-Z., Zhang Y., Liang F.-L., Feng L., (2006), Isolation and characterization of a novel thermophilic Bacillus strain degrading long-chain n-alkanes, Extremophiles, 10, 347-356. Wang Z.D., Fingas M., (1997), Developments in the analysis of petroleum hydrocarbons in oils, petroleum products and oil-spill-related environmental samples by gas chromatography, Journal of Chromatography A, 774, 51-78. Wentzel A., Ellingsen T.E., Kotlar H.K., Zotchev S.B., Throne-Holst M., (2007), Bacterial metabolism of long-chain n-alkanes, Applied Microbiology and Biotechnology, 76, 1209-1221. Yan P., (2006), Alkane-degrading functional bacteria, its cultivation method and application, China Patent and Application, No. CN1789408, 2006–06–21, CN20041081505 20041217, CHENGDU BIOLOGY RES INST OF TH (CN). Yumoto I., Nakamura A., Iwata H., Kojima K., Kusumoto K., Nodasaka Y., Matsuyama H., (2002), Dietzia psychralcaliphila sp nov., a novel, facultatively psychrophilic alkaliphile that grows on hydrocarbons, International Journal of Systematic and Evolutionary Microbiology, 52, 85-90. Yuste L., Corbella M.E., Turiegano M.J., Karlson U., Puyet A., Rojo F., (2000), Characterization of bacterial strains able to grow on high molecular mass residues from crude oil processing, Fems Microbiology Ecology, 32, 69-75. Zarilla K.A., Perry J.J., (1984), Thermoleophilum album gen. nov. and sp. nov., a bacterium obligate for thermophily and normal-alkane substrates, Archives of Microbiology, 137, 286-290. Zhang D., Fakhrullin R.F., Özmen M., Wang H., Wang J., Paunov V.N., Li G., Huang W.E., (2011), Functionalization of whole-cell bacterial reporters with magnetic nanoparticles, Microbial Biotechnology, 4, 89-97. Zhang D., He Y., Wang Y., Wang H., Wu L., Aries E., Huang W.E., (2012), Whole-cell bacterial bioreporter for actively searching and sensing of alkanes and oil spills, Microbial Biotechnology, 5, 87-97. Zhang D., Ding A., Cui S., Hu C., Thornton S.F., Dou J., Sun Y., Huang W.E., (2013), Whole cell bioreporter application for rapid detection and evaluation of crude oil spill in seawater caused by Dalian oil tank explosion, Water Research, 47, 1191-1200. Zhang H.M., Kallimanis A., Koukkou A.I., Drainas C., (2004), Isolation and characterization of novel bacteria degrading polycyclic aromatic hydrocarbons from polluted Greek soils, Applied Microbiology and Biotechnology, 65, 124-131.


July 2013, Vol.12, No. 7, 1359-1365



CONTROL MODE OF RURAL NONPOINT SOURCE POLLUTION IN TAI LAKE BASIN, CHINA Yimin Zhang, Yuexiang Gao, Jingcheng Duan, Yang Yang, Chuang Zhou, Houhu Zhang Nanjing Institute of Environment of Science of Ministry Environmental of Protection, Nanjing 210042, P.R. China

Abstract Eutrophication in Tai Lake Basin is becoming severe in recent years. Although extensive efforts have been made, seeking an alternative cost-effective approach to control water pollution in the area is still an urgent priority. In this paper, the reasons led to the poor control of rural nonpoint source pollution in Tai Lake Basin area were discussed. Thereafter, the paper proposed the ways and means to reduce rural nonpoint source pollutants by a case study in Fenshui village along Caoqiao River of Tai Lake Basin area. A pollution control code including sewage treatment, garbage disposal, farmland runoff pollution control and branch water quality improvement, was introduced. This provides referable measures and methods to solve the pollution problem in Tai Lake Basin. Key words: nonpoint pollution, pollution control mode, Tai lake basin, village Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Tai Lake Basin is placed in one of the most developed and most densely populated areas in South China (Sun and Huang, 1993; Wang et al., 2009). It is also the third largest freshwater lake in China (Qin et al., 2007). In recent years, Tai Lake water pollution has become increasingly serious, especially the eutrophication. Accordingly, algal bloom prevention becomes the focus of water pollution control in Tai Lake area. Reducing the nitrogen and phosphorus loads into the lake is the most appropriate way to diminish this phenomenon. Since the 1970s, the state and the various kinds of enterprises have invested tens of billions of funds for point source pollution control. However, the pollution situation seems not to be fully controlled and there is very little chance to reduce eutrophication (Wang et al., 2009). The reason is the significant increase of industrial activities around the basin area in line with the economic and social development. Another reason was the failure to 

properly control of rural nonpoint source of pollution (Gao et al., 2002; Quan and Yan, 2002; Wang et al., 2009). Rural nonpoint pollution sources are characterized by a large sphere of influence, multiple factors, complicated way, and difficulty in strength quantitative assessment. The pollution comes from many agricultural activities, such as planting, breeding, aquaculture, and rural life (Robert and Cheverry, 1996; Taboada-Gonzalez et al., 2011). These include direct discharge of untreated wastewater from livestock and poultry breeding, high-density aquaculture, excessive feeding (Coote and Gregorich, 2000; Marculescu et al., 2012; Robert and Cheverry, 1996). Also irrational use of agrochemicals, excessive fertilizer, irrational recovery and utilization of agricultural solid waste, remnants of agricultural film in soil and river, and untreated crop straw are other pollution sources. (Patoine and Simoneau, 2002). Wherein, the untreated crop straws are littered along the shore and piled up together with domestic


Zhang et al./Environmental Engineering and Management Journal 12 (2013), 7, 1359-1365

waste. Plenty of percolates are discharged into water bodies or directly into the nearby river during raining (Ndour et al., 2012; Qu et al., 2001; Zhou and Zhu, 2003). Most rural nonpoint source pollutions are accompanied by rainfall runoff. A large number of pollutants enter the water together with surface runoff, without being effectively controlled (Chen et al., 2003). According to the National Environmental Quality Status Report 2008 released by China National Environmental Monitoring Center, Tai Lake was heavily polluted and moderately eutrophic. From 1989 to 2008, permanganate index and total nitrogen of the lake increased, with little change in total phosphorus. From April to September, 2008, sporadic algal blooms and localized algal blooms alternated in the lake, indicating the series situation of pollution. At present, the out-of-specification items in lake water quality include total phosphorus (TP) and total nitrogen (TN); for rivers around the lake, the out-of-specification items include TP and ammonia nitrogen. According to the data of “The first national census communiqué of pollution sources” (NBS,2010), chemical oxygen demand (COD), TN, and TP from agricultural sources accounted for 43.7%, 57.2%, and 67.4%, respectively, of national pollution sources, wherein the proportion of nitrogen and phosphorus pollutants was much higher than those from industrial and domestic sources. Such data further indicated the seriousness of rural nonpoint source pollution and the necessity of emission reduction. Aiming at the status of rural nonpoint source pollution in Tai Lake Basin, the paper proposed the ways and means to reduce rural nonpoint source pollutants prior to presenting a case study in a village in Tai Lake Basin. 2. Case study on nonpoint source pollution control in Fenshui village 2.1. Status of rural nonpoint source pollution in Tai Lake Basin In recent years, people’s living standards have been improved, and the sewage treatment facilities have also been improved continuously in Tai Lake area. The amount of rural pollution shows a decreasing trend. With the reduction in cultivated area and agrochemical input, the amount of nutrient input is also decreasing year by year. As the number of livestock and poultry breeding is gradually stabilized, the farming models become intensified and improved, and the collection and centralized processing of pollutants from livestock and poultry breeding become enhanced. In addition, the pollution emissions from livestock and poultry breeding in Tai Lake Basin show a downward. However, among nonpoint source pollutions of the lake, the proportion of livestock and poultry breeding pollution has an increasing trend. Relatively, the proportion of

1360

domestic pollution and farmland pollution is decreased while the proportion of aquiculture pollution is small. From 1985 to 2008, total phosphorus in Tai Lake Basin was generally decreasing year by year, but total nitrogen and COD have an increasing trend in 2008, as shown in Fig. 1. As the population growth becomes gradually stabilized, the living standard becomes higher and sewage treatment facilities become improved. The amount of rural domestic pollution also gradually declines. Also shown in Fig. 1, the load of farmland pollution in Tai Lake Basin was decreasing from 1980s to 2008. Farmland pollution is mainly caused by large amounts of sediment and pollutants washed down by rainfall runoff, resulting in water environment pollution when they enter into receiving water bodies. Farmland pollution is also affected by a variety of natural and geographical factors, as well as the number of surface contaminants and human irregular activities, such as the irrational use of chemical fertilizers and unscientific farming methods. From 1985 to 2008, the livestock and poultry manure emissions in the Basin are firstly increased and then decreased, as shown in Fig. 1. With the stabilization of breeding number and the continuous intensification and improvement of breeding patterns, as well as the strengthening of pollution collection and centralized processing, pollution emissions from livestock and poultry breeding are reduced. But the proportion in nonpoint source pollution maintains high and there is an increasing trend. In the 1990s, "Three Nets" aquiculture (“Three Nets” refer to net enclosure, net cage, and net pen) in large and medium-sized waters was developed rapidly, which benefits water comprehensive development. With the changes in market demand, the strengthened awareness of water ecological protection and intensified aquaculture structural adjustment, “Three Nets” aquiculture is transformed from the previous high-density, high feeding and high-yield general fish rearing to low feeding, low-yield and high-reward special aquatic farming (crabs, shrimps etc.), which is developed towards industrialization. From 1985 to 2008, pollution loads of total phosphorus, total nitrogen, and COD from aquaculture in all regions of the Basin had the overall trend of first increasing and then decreasing, as shown in Fig. 1. Over the years, the output of nonpoint source pollution from aquaculture is closely related to the number, patterns and species of aquaculture. 2.2. Control mode of non-point source pollution in Fenshui Village Tai Lake Basin is a network with plains, waters and rivers. Eutrophication here is becoming increasingly serious in recent years, which is mostly contributed by the inflow river.

Control mode of rural nonpoint source pollution in Tai Lake Basin, China

The village pollution load reduction and water quality improvement are primarily through source control, treatment and interception.

Fig. 1. The change of pollution loads in Tai Lake Basin

Therefore, treating the inflow river is the key of controlling eutrophication in Tai Lake. Caoqiao River is the main inflow river on the upstream region of Tai Lake. There are various villages located along the river. The pollution load from villages accounts for more than 70-80% of the total pollution load into the river. It is necessary to firstly control the pollution from villages, thereby reducing inflow river load and thus improving the water quality. Therefore, we selected Fenshui Village along Caoqiao River for the case study (Fig. 2). The main pollution source of the village is from domestic sewage, domestic garbage and farmland runoff.

Fig. 2. The location of the Fenshui village and the distribution of the projects

The area of Caoqiao River branch and Fenshui Village is 6 km2. A number of technology demonstrations are considered (Fig. 3). These include sewage treatment for the village in water catchment; multi-stage wetland revetment of farmland runoff; biogas project from domestic garbage and living sewage; and ecological interception of river branches, as shown in Table 1.

Fig. 3. Scheme of non-point source pollution control in village Table 1. Technology demonstration Technology demonstration Sewage treatment for the village in water catchment Multi-stage wetland revetment of farmland runoff Biogas project from domestic garbage and living sewage Ecological interception of river branches

Scale 20m3/d 2 sets, control farmland area within 1km 2 3 sets 3000m

1361


2.2.1. Sewage treatment Based on the local situation of the demonstration area, pipe network was used to collect decentralized rural residents’ sewage. The sewage is treated by centralized process in combination of new biochemical pools, efficient biological filters and constructed wetlands. The sewage treatment project consists of a rural sewage collection system; biochemical pools for microbial immobilization; efficient biological filters; and constructed wetland system (Fig. 4). The sewage collection system involves retrofitting 28 rural residents sanitation facility (toilet, patio and drainage); the construction of septic tanks, concrete roads; pipe laying, as well as drop manholes etc. The sewage treatment capacity is 20m3/d. The size of efficient biological filter unit is 4m×5m×0.8m. The front of the filter is filled with biological sponge. A Muti-Soil-Layering system with the size of 4m×5m×1.1m was built in parallel (Fig. 4).

Pipe collection

Adjusting tank

Efficient biofilter

Constructed wetland

Br anches

MRLsystem

Fig. 4. Diffused village sewage treatment process in the catshment area

The mixing layer materials are the ordinary sand mixed with different proportion of sawdust, charcoal and iron as 8:1:0.5:0.5. The draining layer is made up by zeolite of 40-60mm and the surface load is 0.5 m3/ m2.d (Zhou et al, 2010). The constructed wetland with the size of 5m×6m×0.8m (Huang, et al., 2011) was built as the final unit of the sewage treatment system. Inside the wetland, 20cm of cobble, 40cm of rubble and 20cm of gravel are in turn filled upward in the wetland. The canna was then planted at the top. By operating the treatment system, the removal rates of COD, NH3-N, TN and TP are, respectively, 90-96%, 77-97%, 40-62%, and 90-96%, which meets the requirement of Class A standard of GB18918-2002. 2.2.2. Multi-stage wetland revetment to treat farmland runoff Wetlands are effective at trapping and removing N and P by substrate adsorption, microorganisms decomposition and helophyte vegetation uptake (Raisin and Mitchell, 1995; Jordan et al., 2003; Kang et al., 2002). Two sets of multistage wetland revetment are built, covering 1km2 of farmland area. Solid waste is screened and modified to produce new phosphorus filling for wetland. The wetlands were built in a trapezoidal structure for dropping water, aeration and oxygenation (Fig. 5). The first stage is used to regulate the initial farmland runoff of high concentration while the late runoff is discharged through overflow. The

1362

confluence of the treated runoff then flows into the river. Aeration is realized by the difference in elevation between the riparian farmland and river surface. The revetment can also hold back the initial water flow with high pollution load (Fig. 5). The removal rate of multi-stage wetland revetment is 4060% for COD, 15-25% for TN, 15-15% for TP and 30-50% for SS. 2.2.3. Biogas project from domestic garbage and living sewage The biogas project was conducted by constructing three sets of biogas pools (actually three anaerobic fermentation reactors) for domestic garbage and living sewage treatment while the biogas was generated. The biogas pool uses perishable garbage from rural life as the aggregate while sewage recharge provides nutrients and regulates water. Both garbage and sewage then generate biogas through anaerobic fermentation. Therefore, the waste is treated and fertilizer is produced (Fig. 6). The composition of biogas pool is mainly domestic garbage. Manure and straw are complementary. The sewage addition (via gravity) strengthens the production of biogas and accelerates the maturity. Thus, the goal of "treating rural wastes with rural wastes" is realized and energy regeneration and emission reduction are achieved. Each pool can supply heating for a family of five. The size of biogas is 10m3. The reduction of organics (in terms of carbon), TN and TP is 2-3 t per house per year, 0.40.6 t per house per year and 20-30kg per house per year, respectively (Zhang et al, 2010 ). 2.2.4. Ecological interception The project of ecological interception (Fig. 7) was conducted along the river bank of 3000m by gully bank finishing, sludge dredging, washland regulation, micro-topography regulating, and ecological fences (along the riverside), as well as ecological floating beds (800units 2m×2m) (Jin, et al, 2011). After the construction, the removal rates of 28.6% for COD, 85.7% for TN, 31.8% for TP were achieved. The pollutants flowing to Caoqiao River can then be further reduced. 2.2.5. Water quality improvement via the integrated projects The integrated water environment comprehensive improvement projects include agricultural cleaner production, treatment and utilization of waste from livestock and poultry breeding, aquiculture cleaner production, pollution control of scattered peasant households, channelization, ecological restoration. Additionally, by the support from the local Yixing Municipal Government, other efforts (such as the education on public environmental awareness) have also been made. The pollution load of TN, TP and COD from Fenshui Village into Caoqiao River


was reduced by more than 35% (Table 2). The water environmental quality around the village is significantly improved. The river water quality has been improved to a high level (Table 3). 3. Nonpoint source pollution control measures for villages 3.1. Source control and management Rural nonpoint source pollution control should be integrated efforts, which include the source control, process control of migration and transformation, as well as the end control.

This brings the rural pollutants being treated and utilized locally. In addition, consideration should be made to combine the treatment solution with the adjustment of agricultural structure and the Best Management Practices (BMP). For the source control, it is necessary to improve planting structure and to control the amount of rural domestic sewage, domestic garbage, pesticides and fertilizer in farmland. For the process control, retention and purification are necessary during the migration of contaminants to surface water bodies (Zhang, 2003; Zhang. et al., 2006).

Fig. 5. Farmland/surface runoff wetland treatment process

Biogass Difused domestic garbage Domestic sewage manure

Garbage mathane tank

Methane liquid Methane fertilizer

C transfer to biogass energy utilization

N,P turn into fertilizer back to farmland after ferment

Fig. 6. Biogas project from domestic garbage and living sewage

Fig. 7. Eco-intercepting engineering in Miaojianbang branch

1363


Table 2. Changes of pollutants loading in Fenshui village Pollution load

CODcr

TP

TN

NH4-N

Before the project (t/a) After the project (t/a)

12-15

0.4-0.5

4.5-5.1

3.0-3.5

7.0-10 35%

0.1-0.3 36%

2.5-3.0 38%

1.8-2.2 35%

Reduction (％)

Table 3. Status of water quality in the Caoqiao River at Fenshui village section Concentration Prior to the project (mg/L) After the project (mg/L) Removal rate (％)

CODcr 28.77 (Ⅳ*)

TP 0.36 (V)

18.31(III)

0.22 (IV)

36.4

38.9

TN 9.49 (inferior toV) 3.89 (inferior toV) 59

NH4-N2.1 (inferior toV) 1.11(IV) 47.1

Environmental Quality Standards for Surface Water (GB3838-2002) (Chinese State Environmental Protection Administration, 2002)

Meanwhile, it is also beneficial to consider the waste disposal and utilization, trying to turn waste into treasure, e.g. straw and other solid waste being used to treat domestic sewage (Zhang et al., 2009). For the polluted water bodies in rural areas, it is obliged to treat and recover the polluted water in ecological way. For the whole processes, nonpoint source management should also be strengthened. The overall efforts regarding the nonpoint source pollution control should be extended to the following technical and management strategies: 1) developing green agriculture, ecological agriculture and organic agriculture; 2) adjusting and optimising the planting structure; 3) carrying out quality control of pollution-free agricultural production; 4) promoting cleaner production technology of agriculture; 5) reducing the application amount of chemical nitrogen fertilizers’ and agricultural chemicals; 6) constructing organic agriculture ecosystem; 7) restoring and enhancing the ecological functions of the areas around Tai Lake by building ecological barriers; 8) prohibiting the use of agricultural chemicals from the source in accordance with organic farming standards and production methods; 9) guiding the farmers to grow green manure, strengthening pest monitoring and forecasting, promoting bio-pesticides and higheffective low-poisonous and less-residual pesticides. 3.2. Technology-driven emission reduction Considering the economic development level, technology level and management level, rural areas are not suitable for the urban pattern. It is wiser to adopt measures suitable for local conditions. The fundamental solution to the problem of rural pollutants should be based on agricultural production combined with waste recycling and reuse, which can not only reduce waste emissions, but also achieve resource recovery and utilization. Taking advantage of village decentralized wastewater treatment, rural garbage and solid waste disposal technology (Zhang et al., 2009), the waste 1364

from rural natural villages and surface runoff pollutants should be significantly reduced. In the application of these technologies, pollutants from agriculture and rural life will enter the ecological cycle for recycling and reuse, being treated and utilized locally, thereby reducing the pollutant emissions from village (Zhu et al., 2005). Overall, building an ecocycling mode of nonpoint source pollution control is an important way to solve rural problems. The recommended pollution control measures and interrelations of the different control measures for sustainable development can been see in Fig. 8.

Fig. 8. Control mode of nonpoint source in the Fushui village

4. Conclusions Rural nonpoint source pollution in Tai Lake Basin accounts for an increasing proportion of the overall pollution loading in the area. But the pollution abatement is difficult due to wide range of sources, large amount and difficulty in strength quantitative assessment. This paper proposed a pollution control mode targeting at pollutant recycling and reuse, which has been adapted to local conditions and based on agricultural production and living style of the


residents. The mode could recycle and reuse the “wastes/pollutants” locally and ecologically. This not only reduces the waste emissions, but also achieves resource recovery and utilization, thereby reducing the loading of rural nonpoint source pollution. A case study in a village has demonstrated that the application of the control mode is very beneficial to solve the problem of village pollution and improve the rural water environment in Tai Lake Basin and other similar areas Acknowledgements This study was financially supported by Major Science and Technology Program for Water Pollution Control and Treatment of Chinese Government (2008ZX07101-007).

References Chen X.M., Shen Q.R., Pan G.X., (2003), Characteristics of nitrate horizontal transport in a paddy field of the Tai Lake region, China, Chemosphere, 50,703-706. Cheverry M.C., (1996), Les ressources mondiales en eau et en sols: unelimitation pour l’avenir, Cahiers Agricultures, 5, 43-248. Gao C., Zhu J.G., Dou Y.J., (2002), Contribution of agricultural non-point source pollution to water quality deterioration in Tai lake watershed: recent trends and research priorities, Resource and Environment in the Yangtze Basin (in Chinese),11, 260-263. Huang Z.S., Wei D.Y., Li J., (2011), Imobilization of efficient contact oxidation for treatment of rural domestic sewage on microbe, Journal of Lanzhou Jiaotong University, 5, 259-262. Jin L.D., Ao H.Y., Liu J.T., Li X.P., (2011), Simulation study on in situ purification of lake inlets with artificial floating bed, Journal of Lake Science, 23, 708-718. Jordan T.E., Whigham D.F., Hofmockel K.H., Pittek M.A., (2003), Nutrient and sediment removal by a restored wetland receiving agricultural runoff, Journal of Environmental Quality, 32, 1534-1547. Kang S., Kang H., Ko D., Lee D., (2002), Nitrgen removal from a riverine wetland: a field survey and simulation study of Phragmites japonica, Ecological Engineering, 18, 467-475. Marculescu C., Stan C., Badea A., (2012), Energetic potential assessment of poultry waste processing industry, Environmental Engineering and Management Journal, 11, 1567-1572. NBS, (2010), The first national census communiqué of pollution sources, National Bureau of Statistics of China, online at: http://www.stats.gov.cn/tjgb/qttjgb/qgqttjgb/t2010021 1_402621161.htm. Ndour P., Kane A., Fall A.N., Lazar G., (2012), Rainfall variability in Senegalese rural area: implications on water resources in the borough of Fimela,

Environmental Engineering and Management Journal, 11, 2239-2247. Patoine M., Simoneau M., (2002), Impacts de l’agriculture intensive sur la qualite′ del’eau des rivie` res du Que′ bec, Vecteur Environnement, 35, 61-66. Qin B.Q., Xu P.Z., Wu Q.L., Luo L.C., Zhang Y.L., (2007), Environmental issues of Lake Tai, China, Hydrobiologia, 581, 3-14. Qu W.C., Dickman M., Wang S.M., (2001), Multivariate analysis of heavy metal and nutrient concentrations in Robert, sediments of Lake Tai, China, Hydrobiologia, 450, 83-89. Quan W.M., Yan L.J., (2002), Effects of agricultural nonpoin source pollution on eutrophication of water body and its control measure, Acta Ecologica Sinica, 22, 291-299. Raisin, G.W., Mitchell D.S., (1995), The use of wetlands for the control of non-point source pollution, Water Science and Technology, 32, 177-186. Sun S.C., Huang Y.P., (1993), Tahui Lake (in Chinese), Ocean Press, Bejing, China. Taboada-Gonzalez P., Aguilar-Virgen Q., Ojeda-Benitez S., Armijo C., (2011), Waste characterization and waste management perception in rural communities in Mexico: a case study, Environmental Engineering and Management Journal, 10, 1751-1759. Wang L., Cai Y.L., Fang L.Y., (2009), Pollution in Tai Lake, China: causal chain and policy options analyses, Frontier Earth Science, China, 3, 437-444. Zhang H.H., Zhang Y.M., (2009), Current situation and disposal technology of rural domestic waste in Tai Lake basin, Environmental Sanitation Engineering (in Chinese), 17, 9-11. Zhang H.H., Hu Y., Zhang Y.M., Tian J.S. (2010), Coanaerobic digestion of domestic solid waste and waste water for the decentralized rural household in Taihu Lake Basin, Journal of Ecology and Rural Environment, 26A, 19-23. Zhang Y.M, Zhang Y.C., Zuo Y.H., (2003), Disscussion on application of pre-dam in the nonpoint pollution control of Lake Tai basin, Environmental Pollution and Control, 25, 342-344. Zhang Y.C., Zhang Y.M., Hu M.C., Zhang L.J., Tian M., Tang X.Y., (2006), Studies on front damming technology for NPS pollution control of river network in plain areas, China Water Resources, 17, 14-18. Zhou J., Chen X., Zhang Y.M., Zhang Y.C., Zhou J.R., (2010), Impact of materials in Multi-Soil-Layer (MSL) system on sewage treatment, Journal of Ecology and Rural Environment, 26A, 14-18. Zhou Q.X., Zhu Y.M., (2003), Potential pollution and recommended critical levels of phosphorus in paddy soils of the southern Lake Tai area, China, Geoderma, 115, 45-54. Zhu Y.L., Wu W.L., Huo M., (2005), Ecological countrythe ideal development mode of country in the future, Ecological Economy, 1, 1-3.

1365


INTEGRATED SYSTEM FOR REDUCING ENVIRONMENTAL AND HUMAN-RELATED IMPACTS AND RISKS IN THE WATER USE CYCLE (WATUSER) RESEARCH GRANT 60/2012, PN-II-PT-PCCA-2011 (NATIONAL COLLABORATIVE APPLIED RESEARCH PROJECT), 2012-2015, Project value: 534482 Euro The main objective of the WATUSER project is to develop and implement an integrated system of innovative technologies and management instruments for reducing environmental impacts and associated human health risks caused by water quality aspects in the entire water use cycle: water abstraction, treatment, distribution, use, wastewater collection, wastewater treatment and discharge/reuse. The project implementation consortium includes 2 prestigious Romanian Universities: “Gheorghe Asachi” Technical University of Iasi, which acts as coordinator and Politehnica University of Timisoara, together with 2 important partners from the regional water operators network: SC APAVITAL SA from Iasi and SC AQUATIM SA from Timisoara. For the successful achievement of the main goal, 2 research directions are proposed for the development of:  innovative technologies for water and wastewater treatment in order to respond to specific water quality problems;  a coherent framework of innovative assessment and evaluation instruments that will enable the identification and abatement of environmental and human-related impacts and risks over the water use cycle.

Environmental and human-related impacts & risks and their connection to the water use cycle, within the WATUSER Project The derived specific objectives proposed for the whole water use cycle for the two regional levels of the project (Iasi and Timis Counties) are: 1. Development of specific instruments for the identification, quantification and control of environmental impacts and risks, over the water use cycle, applied to regional water operators; 2. Development of the capacity of collaboration and knowledge transfer between the universities and the regional water operators in Iasi and Timis counties for the control of the environmental impacts and human health risks in the water use cycle; 3. Development of the research and institutional capacities of the universities and regional water operators in Iasi and

Timis counties for facilitation of further cooperation at national and international scale; 4. Development of capacities and competitiveness of Romanian researchers and staff of regional water operators, as well as of the national partnerships contributing to environmental sustainability (protection, conservation of water resources, control of the environmental impacts and human health related risks); 5. Dissemination of relevant results of the project to the scientific community through publication in peer reviewed international journals, ISI ranked, participation in international conferences, workshops, trainings/research stages, as well as to interested stakeholders (industrial agents, water authorities, waterworks companies, agriculture and services, EPAs, local and regional development agencies and authorities, NGOs and societal organizations). The major novel project features are: a) Integration: the environmental, technical, operational and management problems are approached within the water use cycle in an integrated and coherent manner. The integration perspectives of the project refer to multiple aspects: firstly, to the water related problems on the entire water use cycle at the level of regional water operators, taking into consideration the inter-relationships in the complex natural environment–technological anthropogenic system, as well as the spatial-temporal variability of water quality and availability; secondly, the framework of objectives and activities integrates assessment studies with technological development and testing at pilot scale which enables the improvement of regional water operators management as well as the evaluation of the development perspectives of water and wastewater services for small communities. b) Multidisciplinarity: the project activities address problems in multiple science areas related to water resources management: environmental management (water resources management, impact and risk assessment, water footprinting, life cycle assessment, environmental performance indicators) and engineering sciences: chemical and environmental engineering (electrochemical processes, advanced oxidation processes, membrane processes), computer engineering (integrated monitoring system for water related impacts and risk surveys design and implementation). c) Originality, novelty & innovation: the project develops an integrated system of management instrument and technological developments to tackle water related impacts and risks on the whole water use cycle at regional level. To our knowledge, the scale of the project consortium, as well as the complexity of the activities within the WATUSER project represents a premiere for both the Romanian academia and the regional water operators.

Project director: Prof.dr.eng. Carmen Teodosiu Tel.: +40 232 237594; E-mail:[email protected]


July 2013, Vol.12, No. 7, 1367-1373



ENHANCED NITROGEN REMOVAL FROM REJECT WATER OF MUNICIPAL WASTEWATER TREATMENT PLANTS USING A NOVEL EXCESS ACTIVATED SLUDGE (EAS) BASED NITRIFICATION AND DENITRIFICATION PROCESS Yongzhe Yang, Xiaoxia Yang, Wei Li, Xinchao Guo School of Environmental and Municipal Engineering, Xi’an University of Architecture and Technology Xi’an 710055, P.R. China

Abstract In this study, a novel excess activated sludge (EAS) based nitrification and denitrification process was developed using EAS to provide nitrifiers and denitrifers for nitrogen removal from the reject water of a municipal wastewater treatment plant (MWWTP). This can achieve a significant reduction in the nitrogen (N) loading in wastewater treatment plants. The results show that the ammoniacal-nitrogen (NH4+-N) and total nitrogen (TN) removal efficiencies were in the range from 74.7% to 93.1% and 77.4% to 87.9%, respectively. The TN removal efficiency was significantly improved when the step-feed strategy was employed. Therefore, from the technical perspective, the EAS based reject water treatment process can potentially be an alternative pathway for N removal from the reject water. However, for its full-scale application, integrated cost-effective analysis of process capabilities and/or its potential effects on dewaterability of EAS should be considered. Key words: denitrification, excess activated sludge, nitrification, reject water treatment Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction In municipal wastewater treatment plants (MWWTP) that employ biological nitrogen removal (BNR) process for nitrogen (N) removal from wastewater, N is removed by transferring from the aqueous phase into bio-cells and gas phases. Approximately 70% of the influent N is removed through nitrification and denitrification and 20% of the influent N is incorporated in the activated sludge (Henze et al., 2009; Yetilmezsoy et al., 2011). The removal of the excess activated sludge (EAS) from the BNR process can therefore lead to the ultimate removal of bio-fixed N from municipal wastewater. However, during the ensuing anaerobic digestion process for EAS stabilization, N is released from bio-cells into the aqueous phase (Pitman, 1999). 

In a MWWTP the N-enriched filtrate, which is generated in the anaerobic digested sludge dewatering process, termed as reject water (used henceforth in this study) is recycled back to the main wastewater stream for treatment due to the fact that its quality is usually far below the discharge standard of MWWTP. Commonly, the main wastewater treatment process the recycling of reject water leads to 10-30% of the total influent ammoniacal-nitrogen (NH4+-N) loading in. As a result, the effluent quality and N removal efficiency can be negatively affected, particularly when the necessary dissolved oxygen (DO) and sludge retention time (SRT) for nitrification, readily biodegradable chemical oxygen demand (rbCOD) and anoxic retention time for denitrification in the main wastewater treatment process are insufficient (Tchobanoglous et al., 2003).

Author to whom all correspondence should be addressed: E-mail: [email protected]; Phone: +86 13087503299; Fax: +86 29 82202729

Yang et al./Environmental Engineering and Management Journal 12 (2013), 7, 1367-1373

In order to remove NH4+-N from the reject water, the recycle stream can be treated separately with physical processes (such as stripping process) or chemical processes (such as magnesium ammonium phosphate (MAP) precipitation process) (Jenkins et al., 1971). This will accordingly increase the operational cost, the use of chemicals and/or sludge production. The biological processes (SHARON®, Anaerobic ammonia oxidation (ANAMMOX), CANON® and bio-augmentation batch enhanced (BABE) process) are more favourable than the physical-chemical processes (Berends et al., 2005; Dosta et al., 2007). The biological nitrogen removal processes have been evaluated for the reject water treatment. However, these approaches will further increase the difficulty in operation and maintenance of treatment systems, particularly in the growth, accumulation and/or selection of ammonia-oxidizing bacteria (AOB), nitrite-oxidizing bacteria (NOB) and/or the ANAMMOX bacteria (Ghyoot et al., 1999; Hwang et al., 2000). While the limited denitrification and the resultant NO2 and NO gas, which are either toxic or strong greenhouse gases, may have detrimental effect on the applicability of such biological process for reject water treatment (Henze et al., 2009; Xu et al., 2011). On the other hand, the EAS generated from the BNR process contains high level of AOB and NOB that may be able to convert ammonia and/or nitrite. The EAS also contains intracellular carbon source, such as glycogen and/or polyhydroxyalkanoates (PHAs), which can be a potential electron donor for denitrification (Zafiriadis et al., 2011; Zeng et al., 2011). From the literature, it seems that there is no any report on the utilization of EAS for NH4+-N removal from rejected water by using the capacity of excess sludge in the nitrification and denitrification. This can be an alternative pathway for reject water treatment and can ease the pressure and even solve the problems of N feedback through reject water recycling. Therefore, the main objectives of this study are (1) to demonstrate nitrification performance of utilizing EAS as nitrifiers for NH4+-N removal from reject water and (2) to investigate the feasibility of using EAS for the ensuing denitrification for nitrate and/or nitrite removal for the treated reject water. 2. Materials and methods 2.1. Materials 2.1.1. Reject water collection The reject water was collected periodically from the holding tank of NO.1 MWWTP in Xi’an city, P.R. China. The MWWTP employs an anaerobic/anoxic/aerobic (A2/O) process (Hydraulic Retention Time (HRT): 18hrs, SRT: 20-25 days) for municipal wastewater treatment and an anaerobic digester (HRT=30-33 days) for waste sludge stabilization.

1368

The reject water is the supernatant from the belt press (anaerobic digested sludge dewatering unit), which accounts for approximately 1% or less of the influent flow rate of the MWWTP. The average total Kjeldahl nitrogen (TKN) and NH4+-N in the raw reject water were 786.6 ± 70.9 mg-N/L and 751.7 ± 71.1 mg-N/L, respectively. The average alkalinity of the raw reject water was 1956.5 ± 187.3 mg-CaCO3 /L. 2.1.2. Excess activated sludge collection EAS was obtained from the holding tank where the secondary settled activated sludge were collected and pumped to the main wastewater treatment stream or the waste sludge treatment process. The flow rate of the EAS accounts for 1.53% of the entire influent of the MWWTP. The suspended solid (SS) in the EAS was ranged from 6506.1 to 9023.0 mg/L with an average of 7109.3 mg/L. The alkalinity and NH4+-N of the EAS ranged from 183.4 to 236.7 mg-CaCO3/L and 5.9 to 18.9 mg-N/L, respectively. The main characteristics of the raw reject water and the EAS are shown in Table 1. 2.2. Methods 2.2.1 Laboratory scale EAS based SBR set-up and operation This study was carried out on a 12 L laboratory scale sequencing batch reactor (SBR) using the EAS to act as the nitrifiers and/or denitrifers. The EAS based SBR was operated under a 48-hour aerobic-anoxic cycle, i.e. 0.5 h for filling, 23.5 h for aerobic reaction at dissolved oxygen ranged from 3.0 - 5.0 mg/L, 23.5 h for anoxic reaction (with EAS secondary step-feed) and 0.5 h draw. Schematic description of the EAS based SBR operation and feed strategies are shown in Fig. 1. The reject water and the EAS were mixed (at mix ratios of 15%:85%:0%, 15%:80%:5%, 15%:75%:10% and 15%:65%:20%, respectively (volume calculated according to the volume of the lab-scale SBR). In the aeration stage of the SBR, 15% of the reject water was mixed with 85%, 80%, 75% and 65% of the EAS, respectively. The step-feed strategy was employed to enhance the denitrification performance of the treatment system. This was achieved by further feeding 5%, 10% and 20% of the EAS in the anoxic stage of the SBR according to the initial mixture of the reject water and the EAS at mix ratios of 15%:80%,15%:75% and 15%:65%, respectively. The initial alkalinity of the mixture in the aeration stage of the SBR was maintained in the range of 850 to 950 mg-CaCO3/L using sodium carbonate (0.1 M). 2.2.2. Analysis Samples of the raw reject water, EAS and mixtures of the raw reject water and EAS and the mixed liquor from each stage of the EAS based SBR system were collected for SS measurement.

Enhanced nitrogen removal from reject water of municipal wastewater treatment plants

Table 1. Characteristics of the raw reject water and the EAS Parameter TKN (mg-N/L) NH4+-N (mg-N/L) NO3--N (mg-N/L) NO2--N (mg-N/L) SCOD (mg/L) SS (mg/L) Alkalinity (mg-CaCO3/L)

Reject water (Average ± SD, n=156) 786.6 ± 70.9 751.7 ± 71.1 101.2 ± 23.4 143.6 ± 29.1 1956.5 ± 187.3

EAS (Average ± SD, n=156) 8.9± 3.2 19.5 ± 6.2 2.1 ± 0.7 57.1 ± 18.3 7109.3 ± 666.2 198.5 ± 41.3

Reject water (15%) EAS (85%-80%-75%-65%)

Fill: 0.5h

100% Mixed liquor (to thickening unit) Aeration: 23.5h (Nitrification)

Draw: 0.5h

EAS (0%-5%-10%-20%)

Mix: 23.5h (Non aeration) (Denitrification)

Fig. 1. Schematic description of the EAS based SBR operation and feed strategies

Then the samples were filtered using 0.45 μm filter membrane for soluble chemical oxygen demand (SCOD), TKN, NH4+-N, nitrate-nitrogen (NO3--N) and nitrite-nitrogen (NO2--N) measurement. SCOD, TKN, NH4+-N, NO3--N and NO2--N were analysed using standard methods of 5210 B, 4500Norg B, 4500-NH3 C, 4500-NO3- B and 4500-NO2- B, respectively (APHA, 1992). 3. Results and discussion 3.1. Nitrogen removal performance of the EAS based SBR Fig. 2 shows the removal performance of NH4+-N in the EAS based SBR. The NH4+-N concentration in the raw water of the treatment system, which the reject water was mixed with the EAS at mix ratios of 15% to 85%, 15% to 80%, 15% to 75% and 15% to 65%, ranged from 108.7 to 156.4 mg-N/L, 113.9 to 156.3 mg-N/L, 119.9 to 165.2 mgN/L and 133.9 to 177.5 mg-N/L, respectively. At various mix ratios of the reject water and EAS, the EAS based reactors showed excellent NH4+-N removal performance. The NH4+-N was reduced in the treatment system with effluent NH4+-N concentration ranged from 7.9 to 26.8 mg-N/L, 9.9 to 33.3 mg-N/L, 10.1 to 33.7 mg-N/L and 11.9 to 31.6 mg-N/L, respectively. The EAS based SBR removed

significant amount of NH4+-N from the influent wastewater. The results show that the average NH4+-N removal efficiencies were in the range from 79.7% to 93.1%, 74.7% to 91.6%, 75.3% to 92.1% and 79.6% to 91.2%, respectively. It is well known that Nitrosospira- and Nitrosomonas-related AOB appeared to be prevalent in activated sludge treating sewage (Hiorns et al., 1995; Wagner et al., 1998, Purkhold et al., 2000). According to Whang et al. (2009) and Kim et al. (2011), Nitrospira- and Nitrobacter- like bacteria were the dominant NOB in MWWTP. In this study, the EAS was wasted from the return sludge line of the main wastewater treatment process of the MWWTP where the AOB and NOB were developed in the aerobic zone of the treatment system. Therefore, the proportion and properties of the AOB and NOB in the EAS and the mixed liquor of the aerated zone is the same. This suggests that EAS can provide the nitrifiers for NH4+-N removal from the reject water. Fig. 3 shows the removal performance of TN in the EAS based SBR. The average TN concentration in the raw water of the treatment system, which the reject water was mixed with the EAS at mix ratios of 15% to 85%, 15% to 80%, 15% to 75% and 15% to 65%, were 153.1mgN/L, 152.5mg-N/L, 163.6mg-N/L and 173.9 mg-N/L, respectively.

1369

Reject water:EAS:EAS (b) 15%:80%:5%

Reject water:EAS:EAS (a) 15%:85%:0%

Reject water:EAS:EAS (d) 15%:65%:20%

Reject water:EAS:EAS (c) 15%:75%:10%

100

200

60

100

+

80

Removal efficiency

150

Influent

50

40 20

0

0

10

20

30

0

10

20

30

0

10

20

30

0

10

20

30

+

Effleunt

NH4 -N removal efficiency (%)

NH4 -N concentration (mg-N/L)


40

Operation (number of samples)

Reject water:EAS:EAS (b) 15%:80%:5%

Reject water:EAS:EAS (a) 15%:85%:0% Removal efficiency

250

Reject water:EAS:EAS (d) 15%:65%:20%

Reject water:EAS:EAS (c) 15%:75%:10%

100

200

80

150

60

100

Influent Effleunt

40

50 0

20 0

10

20

30

0

10

20

30

0

10

20

30

0

10

20

30

TN removal efficiency (%)

TN concentration (mg-N/L)

Fig. 2. NH4+-N removal performance of the EAS based SBR under different feed strategies

40

Operation (number of samples) Fig. 3. TN removal performance of the EAS based SBR under different feed strategies

The average effluent TN concentrations were 34.3 mg-N/L, 34.4 mg-N/L, 33.0 mg-N/L and 21.1 mg-N/L, respectively. The TN removal efficiency was improved when the step-feed strategies were employed. The results show that the average TN removal efficiencies were 77.4%, 77.5%, 79.7% and 87.9%, respectively. The results indicate that the more the EAS fed to the anoxic stage of the EAS based SBR, the higher the TN removal efficiency was obtained, indicating that the more carbon source can be utilized for denitrification. However, the SCOD in the EAS was relatively low (see Table 1) as compared to the organic carbon for denitrification. This reveals that intracellular carbon sources that contained in the EAS can potentially act as the electron donor for denitrification (Wu and Rodgers, 2010). 3.2. Characteristics of nitrification and denitrification under different feed strategies Fig. 4 shows the characteristics of nitrification and denitrification under different operation and feed strategies using mean value and standard deviation (S.D) of NH4+-N, NO3--N and NO2--N, which were calculated based on 6 cycles of the EAS based SBR. The results show that the mean NH4+-N concentration was decreased from 128.6 ± 7.7 mg1370

N/L to 14.0 ± 2.4 mg-N/L in the aerobic stage of the treatment system. In the aerobic stage, the reduction in NH4+-N concentration occurred simultaneously with increases in concentrations of NO3--N (0.3 ± 0.2 to 107.4 ± 6.3 mg-N/L) and NO2--N (0.02 ± 0.03 to 1.4 ± 0.9 mg-N/L) suggests that the EAS can be utilized as ammonia oxidizing biomass, indicating that nitrification is the dominating mechanism and pathway for NH4+-N removal. In the anoxic stage, various amount of the EAS was fed and the effluent NO3--N were 18.3 ± 2.6 mg-N/L, 16.7 ± 4.3 mg-N/L, 13.8 ± 2.1 mg-N/L and 0.99 ± 2.1 mg-N/L for various step-feed strategies of 15%:85%:0%, 15%:80%:5%, 15%:75%:10% and 15%:65%:20% (v/v, reject water: EAS:EAS). The step-feed strategy of 15%:65%:20% (reject water: EAS:EAS) presents the highest denitrification ability. In the typical BNR process, depletion of the cellular stored glycogen and/or PHAs was observed during the denitrification reaction (Winkler et al., 2011). As the EAS used in this study was the same source as the mixed liquor of the BNR process, it is reasonable to suggest that the resultant high nitrate removal ability is as a result of denitrification using the intracellular carbon source in the EAS and/or the step-feed strategy supplies the carbon necessary for NO3--N removal.


ensuing anoxic unit (HTR=24h), 20% of the EAS was fed to achieve better denitrification performance. If the EAS based nitrification-denitrification process was applied, according to the experimental result from this study, the TN in the reject water can be reduced from 786.6 mg-N/L to 21.0 mg-N/L. This gives a TN removal efficiency of 83.9%. However, the drawbacks of these proposals lie in (1) the potential increases of operation costs and consume more space, (2) the extra capital and aerobic and/or anoxic tank for the reject water treatment and (3) its potential effects on dewaterability of EAS. It should be noted that emphasis has always been located on the disadvantages that may occur, rather than on the potential advantages that such EAS based nitrification-denitrification process offers. The proposed processes have the potential benefit of NH4+-N and/or TN reduction in reject water, which reduces the nitrogen loading in the wastewater treatment process when the reject water is recycled. From the technical point-of-view, this strategy is practicable. The choice of the treatment process for reducing ammoniacal-nitrogen loading will be always highly depend on the specific limitation in each MWWTP. For example, if the aeration capacity in the MWWTP is insufficient, then the EAS based nitrification for ammonium removal will be the proper process to be selected (as shown in Fig. 5(a)). While if the rbCOD is insufficient for denitrification in the main wastewater treatment stream, then only the EAS based nitrificationdenitrification employing step-feed strategy would be the proper process to be chosen (as shown in Fig. 5(b)).

3.3. Case analysis Fig. 5 (a, b) shows the potential application of the EAS based nitrification-denitrification process on NH4+-N and/or TN removal from the reject water in MWWTP. Fig. 5(a) proposes integration of the EAS and the reject water in the aerobic unit under a determined mix ratio of 5:1 (EAS : reject water, volume basis) using a hydraulic retention time (HRT) of 24 h. This allows the nitrification to take place and transfer NH4+-N to NO2--N and/or NO3--N, thus eliminating the NH4+-N level in the reject water. Mass balance shows that the NH4+-N in the reject water, which is currently recycled to the beginning of the biological treatment system in the MWWTP, is 751.1 ± 71.1 mg-N/L. If the EAS based nitrification process was applied, according to the experimental result from this study, the computed NH4+-N concentration in the reject water could be reduced from 751.1 ± 71.1 to 18.8± 1.8 mg-N/L, as shown in Fig. 5(a). A significant reduction of NH4+-N loading in the reject water is therefore achieved, which will reduce approximately 85% of the NH4+-N loading. Reduced NH4+-N level in the reject water may help to minimize the effect of the recycled nitrogen on effluent quality of the MWWTP. To achieve better TN removal and to reduce the nitrate-nitrogen and/or nitrite-nitrogen feed-back to the main wastewater stream in the MWWTP, Fig. 5(b) proposed integration of 80% of EAS and the reject water in the aerobic unit under a determined mix ratio of 4:1 (EAS : reject water, volume basis) using a HRT of 24 h to transfer NH4+-N to NO2--N and/or NO3--N. In the (a) R eject w ater:E A S :E A S = 1 5 % :8 5 % :0 %

+

-

-

NH4 , NO3 and NO2 concentration (mg-N/L)

150

A ero b ic stag e

125

(b ) R eject w ater:E A S :E A S = 1 5 % :8 0 % :5 % A ero b ic stag e

A n o x ic stag e

A n o x ic stag e

+

N H 4 -N

100

-

N O 3 -N

75 50

-

N O 2 -N

25 0

(d ) R eject w ater:E A S :E A S = 1 5 % :6 5 % :2 0 %

(C ) R eject w ater:E A S :E A S = 1 5 % :7 5 % :1 0 %

150

A n o x ic stag e

A ero b ic stag e

A ero b ic stag e

A n o x ic stag e

125 100 75 50 25 0

0

6

12

18

24

30

T im e (h o u r)

36

42

48 0

6

12

18

24

30

36

42

48

T im e (h o u r)

Fig. 4. Characteristics of nitrification and denitrification under different feed strategies (Mean value and standard deviation (S.D) of NH4+-N, NO3--N and NO2--N obtained under each feed strategy were calculated based on 6 individual cycles of the EAS based SBR.)

1371


Influent (Flow rate:Q)

Anaerobic/Anoxic/Aerobic Tank

Secondary Clarifier

Effluent

Aerobic Unit (HRT: 24h, Flow rate:3.6%Q) Supernatant with low NH4+-N

Thickening Unit

(Flow rate: Approximately 2.95%Q)

Anaerobic Digester

Reject water (Flow rate:0.6%Q)

Excess activated sludge (Flow rate:3%Q)

Dewatering Unit Dewatered sludge (cake) (Approximately 0.05%Q)

a) Influent (Flow rate:Q)

Anaerobic/Anoxic/Aerobic Tank

Secondary Clarifier

Effluent

(Flow rate:2.4%Q)

Aerobic Unit (HRT: 24h, Flow rate:3%Q) Anoxic Unit (HRT: 24h, Flow rate:3.6%Q)

Supernatant with low TN (Flow rate: Approximately 2.95%Q)

Thickening Unit

Reject water (Flow rate:0.6%Q)

EAS (Flow rate:0.6%Q)

EAS (Flow rate:3%Q) EAS

Anaerobic Digester Dewatering Unit Dewatered sludge (cake) (Approximately 0.05%Q)

b) Fig. 5. Strategic outline of N immobilization from the reject water using the EAS based nitrification-denitrification process: a) EAS based nitrification process; b) EAS based nitrification-denitrification process

4. Conclusions The outcome of this study has shown that the EAS can be successfully used to act as the nitrification and denitrification biomass. The application of the EAS based nitrification and denitrification beneficially enhanced NH4+-N and/or TN removal from the reject water, thereby eliminating the N loading in the wastewater treatment process when the reject water is recycled. The major conclusions are as follows: ● The EAS has demonstrated a considerable ability for NH4+-N removal from the reject water. The dominating pathway for NH4+-N removal is nitrification. The EAS based nitrification process can potentially remove approximately 85% of the NH4+N loading. ● The EAS has also demonstrated excellent denitrification ability. Particularly, the denitrification 1372

performance can be significantly improved when step-feed strategy was employed. According to this study, the EAS based nitrification-denitrification process will achieve a reduction of approximately 83.9% of TN in the recycled reject water. ● The EAS based reject water treatment process was also demonstrated as an alternative pathway of N removal from aqueous phase through nitrification and/or denitrification. Acknowledgements The authors wish to acknowledge financial support obtained from the Key Program of Natural and Science Foundation for Applied Basic Research of Shaanxi Province (Project NO.: 2010JZ008) and Specialized Research Fund for the Doctoral Program of Higher Education (SRFDP Project No.: 20116120110008). Sincere thanks are given to the reviewers for their valuable comments to improve the paper.


References APHA, (1992), Standard methods for the examination of water and wastewater (18th Edition), American Public Health Association, Washington DC. Berends D.H.J.G., Salem S., Roest H.F.V.D., Loosdrecht, M.C.M.V., (2005), Boosting nitrification with the BABE technology, Water Science and Technology, 52, 63-70. Dosta J., Gali A., El-Hadj T.B., Mace S., Mata-Alvarez, J., (2007), Operation and model description of a sequencing batch reactor treating reject water for biological nitrogen removal via nitrite, Bioresource Technology, 98, 2065-2075. Ghyoot W., Vandaele S., Verstraete, W., (1999), Nitrogen removal from sludge reject water with a membraneassisted bioreactor, Water Research, 33, 23-32. Henze M., Loosdrecht M.C.M.V., Ekama G.A., Brdjanovic, D., (2009), Biological Wastewater Treatment: Principles, Modelling and Design, IWA Publishing, London, UK. Hiorns W.D., Hastings R.C., Head I.M., McCarthy A.J., Saunders J.R., Pickup R.W., Hall G.H., (1995), Amplification of 16S ribosomal RNA genes of autotrophic ammonia-oxidizing bacteria demonstrates the ubiquity of nitrosospiras in the environment, Microbiology, 141, 2793-2800. Hwang Y., Yoneyama Y., Noguchi H., (2000), Denitrification characteristics of reject water in upflow biofiltration. Process Biochemistry, 35, 1241-1245. Jenkins D., Ferguson J.F., Menar A.B., (1971), Chemical processes for phosphate removal, Water Research, 5, 369-389. Kim Y.M., Cho H.U., Lee D.S., Park D., Park J.M., (2011), Influence of operational parameters on nitrogen removal efficiency and microbial communities in a full-scale activated sludge process. Water Research, 45, 5785-5795. Pitman A.R., (1999), Management of biological nutrient removal plant sludges - Change the paradigms? Water Research, 33, 1141-1146. Purkhold U., Pommerening-Röser A., Juretschko S., Schmid M.C., Koops H-.P., Wagner M., (2000), Phylogeny of all recognized species of ammonia oxidizers based on comparative 16S rRNA and amoA sequence analysis implications for molecular diversity

surveys, Applied Environmental Microbiology, 66, 5368–5382. Tchobanoglous G., Burton F.L., Stensel H.D., (2003), Wastewater Engineering: Treatment and Reuse (4th Edition), Metcalf & Eddy, Inc., New York. Wagner M., Noguera D.R., Juretschko S., Rath G., Koops H-.P., Schleifer K-.H., (1998), Combining fluorescent in situ hybridization (FISH) with cultivation and mathematical modeling to study population structure and function of ammonia-oxidizing bacteria in activated sludge, Water Science and. Technology, 37, 441-449. Whang L.-M., Chien I.-C., Yuan S.-L., Wu Y.-J. (2009), Nitrifying community structures and nitrification performance of full-scale municipal and swine wastewater treatment plants, Chemosphere, 75, 234242. Winkler M., Coats E.R., Brinkman C.K., (2011), Advancing post-anoxic denitrification for biological nutrient removal, Water Research, 45, 6119-6130. Wu G.X., Rodgers M., (2010), Dynamics and function of intracellular carbohydrate in activated sludge performing enhanced biological phosphorus removal, Biochemical Engineering Journal, 49, 271-276. Xu D.F., Li Y.X., Xu X.H., Zhao X.L., Fang H., (2011), Effect of microbial activity in the rhizosphere of wetland plants on removal of total organic carbon and nitrogen from wastewater, Environmental Engineering and Management Journal, 10, 781-786. Yetilmezsoy K., Turkdogan-Aydinol I., Gunay A., Ozis I., (2011), Post treatment of poultry slaughterhouse wastewater and appraisal of the economic outcome, Environmental Engineering and Management Journal, 10, 1635-1645. Zafiriadis I., Ntougias S., Nikolaidis C., Kapagiannidis A.G., Aivasidis A., (2011), Denitrifying polyphosphate accumulating organisms population and nitrite reductase gene diversity shift in a DEPHANOX-type activated sludge system fed with municipal wastewater, Journal of Bioscience and Bioengineering, 111, 185-192. Zeng W., Li L., Yang Y.Y., Wang X.D., Peng Y.Z., (2011), Denitrifying phosphorus removal and impact of nitrite accumulation on phosphorus removal in a continuous anaerobic-anoxic-aerobic (A2O) process treating domestic wastewater, Enzyme and Microbial Technology, 48, 134-142.

1373

ECOMONDO is one of the prominent EU industrial exhibitions on material recovery and valorization, resource and energy efficiencies, and sustainable industrial growth. It is hosting the exhibition stands of about 1200 major enterprises along with over than 100 scientific/technical workshops and conferences in a 100,000 square meters facility visited by over than 80,000 delegates from industries, academia and governmental institutions from over 40 different countries, mostly from the Mediterranean basins. The major industrial areas covered in terms of policy, RTD and innovation by ECOMONDO exhibition and workshops/conferences are, among others:  waste production reduction, waste collection, fractionation, recycling, exploitation  biomass and biowaste exploitation via integrated biorefinery scheme and bioplastics and biobased products; Green and sustainable chemistry have also key priorities of this area  water resources monitoring and protection; wastewater treatment and valorization with nutrients recovery; marine resources protection and exploitation  sustainable remediation of contaminated sites and marine ecosystems  indoor and outdoor air monitoring and clean up. The major innovative tools, technologies and solutions related to the sustainable management of waste, water, air and energy in our urban living are implemented in a confined area of 6,000 m2 where a pilot scale smart/sustainable city is built up to display the innovation to experts, citizens, students and the others delegates who can see, test, and validate the proposed technologies and innovation products. Then another area of ECOMONDO is devoted to European Eco Innovation in the field of green economy. Here Horizon2020 priorities and relevant innovation funding opportunities will be presented by delegates of the EU commission; the coordinators of the Biobased Industry JTI and those of the PPP on resource efficiency and waste exploitation (BRIDGE), the EIP on water and raw materials, the KIC climate and that on raw materials will present the major objectives and actions of their initiatives. This to provide information on the major EU innovation policies, actions and funding opportunities and catalyze the formation of industry-academia and NGO partnerships interested in applying for the EU funding opportunities offered by the EU commission (Horizon2020 and other EU funding system). Another initiative is dedicated to the promotion of industrial symbiosis and networking, i.e. the creation of partnerships between industries working in the complementary sectors or different countries; in the same frame, dissemination activities associated with the most prominent international and EU RTD and industrial projects are also hosted. Thus, ECOMONDO is a special opportunity for SMEs and industries, which can find a fertile and dynamic international environment for presenting/proposing their products and technologies (via the exhibition area) and participating in scientific technical conferences, meeting new industrial and academic partners, knowing new EU innovation policies, priorities and opportunities and being involved in partnerships applying for public funds for supporting eco-innovation. It is also a special opportunity for the academia and the young scientists in particular, who can attend and present their work in scientific technical conferences, publish their results on a international journal, meet relevant industries, enter the EU innovation and being involved in partnerships applying for grants supporting ecoinnovation. The same is applying for the public institutions, which can be rapidly updated on the new national and EU industrial policies and innovation priorities and the major international players that are promoting them. Thus ECOMONDO is a relevant “Euro-Mediterranean Technology Platform” on Green Economy. Professor Fabio Fava Department of Civil, Chemical, Environmental and Materials Engineering University of Bologna Vice-Chair - Environmental Biotechnologies of the European Federation of Biotechnology (EFB) Chair of the Technical and Scientific CommitteeECOMONDO 2013 - RIMINI 6-9 november 2013


July 2013, Vol.12, No. 7, 1375-1380



AEROBIC SLUDGE GRANULATION FOR PARTIAL NITRIFICATION OF AMMONIA-RICH INORGANIC WASTEWATER An-Jie Li1, Xiao-Yan Li2, Xiang-Chun Quan1, Zhi-Feng Yang1 1

Key Laboratory of Water and Sediment Sciences of Ministry of Education / State Key Joint Laboratory of Environment Simulation and Pollution Control, School of Environment, Beijing Normal University, Beijing 100875, P.R. China 2 Environmental Engineering Research Centre, Department of Civil Engineering, The University of Hong Kong, Pokfulam Road, Hong Kong, P.R. China

Abstract Partial nitrification for the biological nitrogen removal (BNR) via nitrite has recently gained interest for treating high-strength ammonium wastewater with low organic matter content. However, stability of partial nitrification is still a challenge in activated sludge system. In the present study, the technique of aerobic granulation was developed to produce granules for stable nitritation, treating ammonia-rich inorganic influent with 400 mg NH4+-N/L. The morphology, physical properties, bacterial community structure and partial nitrification performance of the sludge were characterized throughout the experiments. The results indicated that aerobic granules could be produced for partial nitrification through selective discharge of small and slow-settling sludge flocs. Sludge granulation help to achieve ammonium oxidation to the level of nitritation, or partial nitrification, other than to complete nitrification. Based on DNA-base molecular analysis, aerobic granulation resulted in an enrichment of AOB and a reduction of nitrite-oxidizing bacteria (NOB) in the granular sludge, which is highly favorable to a stable operation of partial nitrification. Key words: aerobic granulation, biological wastewater treatment, inorganic wastewater, partial nitrification Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Partial nitrification for the biological nitrogen removal (BNR) via nitrite has recently gained interest because of its associated economical savings when compared to total nitrification, especially for treating high-strength ammonium wastewater with low organic matter content (Bautista-Patacsil et al., 2012; Han et al., 2010; Turk and Mavinic, 1987; Van Hulle et al., 2010). The nitrite produced by ammonia-oxidizing bacteria (AOB) in partial nitrification could be readily removed via denitrification by anaerobic ammonium oxidation (anammox) or other similar processes (Ji and Chen, 2010; Ostace et al., 2012; Zhang et al., 2010). A partial nitrification or nitritation (ammonium oxidation to nitrite) could be obtained in activated sludge systems (Jubany et al., 

2009; Yamamoto et al., 2008) or biofilm (Bartroli et al., 2010; Furukawa et al., 2009). However, several studies have reported problems about stability of partial nitrification (Fux et al., 2004; Yun and Kim, 2003) due to accumulation of nitrite-oxidizing bacteria (NOB) easily in the nitrite-enrich condition. Sufficient washout of NOB is fundamental for a stable partial nitrification and the sludge retention time (SRT) is one of the key parameters to achieve it (de Graaff et al., 2010; Jubany et al., 2009). The yield of AOB is higher than that of NOB at room temperature (Blackburne et al., 2007). Therefore, the shorter the SRT, the easier the NOB washout is at room temperature. However, the activated sludge systems generally work at long SRT to promise treatment efficiency due to the low biomass concentration of the system.

Author to whom all correspondence should be addressed: E-mail: [email protected]


Aerobic granulation is a process in which loose sludge flocs are transformed into dense granules. Due to attributes such as a compact structure and fast settling velocity, granular sludge allows a high level of biomass retention, a very short phase of sludge sedimentation and a much higher loading rate in bioreactors (de Kreuk and van Loosdrecht, 2006; Liu and Tay, 2004). Granule production could achieve high biomass concentration leading to a high rate of ammonium oxidation (Tsuneda et al., 2003). However, due to the slow growth rates of nitrifying bacteria, complete granulation is rather difficult for nitrification or partial nitrification, especially for the concentrated ammonia feed with little organic substrates (Belmonte et al., 2009). In a few successful cases reported, a long start-up period of 2 months or longer was needed to achieve sludge granulation (Kim and Seo, 2006; Lopez-Palau et al., 2011; Shi et al., 2010). Selective discharge loose and slow-settling flocks were strategies for the fast start-up of sludge granulation (Li and Li, 2009; Qin et al., 2004). Therefore, effective cultivation strategies need to be developed to accelerate granule formation for reliable granulation and stable partial nitrification in treating ammonia-rich inorganic wastewater. The aims of the experimental study were to develop an effective technique for rapid granulation of AOB sludge for nitritation under high inorganic loading, to investigate the microbial population transformation during the granulation process and to examine the stability of partial nitrification by the granular sludge. 2. Materials and methods

KH2PO4 without any organic substrates added. Filtered clean seawater was added into the wastewater to increase its salinity to 1% to simulate the saline wastewater in Hong Kong and provide other inorganic salts for microbial growth. The wastewater influent contained an NH4+-N concentration of 400 mg N/L and a PO43--P concentration of 40 mg P/L, resulting a volumetric N loading of 0.8 g N/L·d in the reactor. The pH of the mixed liquor in the reactor was controlled at around 7.5 during the experimental period by adding a diluted NaHCO3 (0.1 M) solution automatically. The reactor was operated at room temperature, and the water temperature was 20-22˚C. Discharge of small and slow-settling sludge flocs was conducted at the end of each 12-h cycle from the bioreactor. During the phase of sludge discharge, the sludge was allowed to settle in the column without aeration. The settling period varied from 1 to 5 min depending on the sludge settling property and the targeted amount of sludge to be discharged. The slow settling sludge in the suspension was therefore removed from the bioreactor. The amount of the daily sludge discharge and the sludge concentration in the bioreactor were measured. The amount of daily sludge discharge was adjusted accordingly to maintain a biomass MLVSS concentration of around 2000 mg/L in the bioreactor. After the selective sludge discharge, the remaining sludge suspension was allowed to settle in the column for another 30 min in order to minimize sludge loss in the effluent during effluent withdrawal. The supernatant was then withdrawn from the bioreactor, and the feed solution was added into the bioreactor to restore the original volume of 200 mL.

2.1. Experimental set-up and SBR operation A small column with height of 30 cm and internal diameter of 3.6 cm and a working volume of 200 mL was used as the sequencing batch reactor (SBR) to grow granular sludge for partial nitrification (Fig. 1). The bioreactor was operated in a fixed sequential mode for a 12-hr cycle with 1 min of feeding, 11 hr and 53-57 min of aeration, 1-5 min of sludge settling and 1 min of effluent withdrawal. The volume exchange ratio of the bioreactor was about 80%. The reactor was inoculated with nitrifying activated sludge that had been cultivated in a labscale fermentor (Sartorius Biostat® A plus, Germany). The feed into the fermentor was an NH4Cl solution with a NH4+-N concentration of 200 mg/L and no organic content (Ye and Zhang, 2010). The initial sludge concentration added in the column reactors was around 2000 mg/L in terms of mixedliquor volatile suspended solids (MLVSS). Aeration was supplied from the bottom of bioreactor by an air pump at a flow rate of 8 L/min to keep the DO concentration in the sludge suspension in a range of 2-4 mg/L. The influent to the reactor was a synthetic wastewater prepared with NH4Cl and

1376

Fig. 1. Schematic of the bioreactor

Aerobic sludge granulation for partial nitrification of ammonia-rich inorganic wastewater

2.2. Analytical methods The sludge MLVSS concentration and the sludge volume index after 5 min sedimentation (SVI5) were measured according to the Standard Methods (APHA 2005). The concentrations of nitrogen (e.g. NH4+-N, NO2−-N and NO3−-N) in the discharge suspension from the bioreactor after sludge settling were measured at the end of each 12-h cycle during the experiment. The ammonia, nitrite and nitrate concentrations were measured following the Nesslerizaion method, colorimetric method, and ultraviolet spectrophotometric screening methods, respectively (APHA 2005). The morphology of the sludge flocs and granules was examined under a stereomicroscope (S8APO, Leica, Cambridge, UK) equipped with a digital camera (EC3, Leica, Cambridge, UK). A laser diffraction particle counter (LS13320, Beckman Coulter, Miami, FL, USA) was used to determine the size distribution of the sludge flocs and granules. Accordingly, the volume-based mean size of the sludge in a sample was calculated. 2.3. Settling velocity and nitrification kinetics of granular sludge The settling velocity was measured for individual granules by recording the time for the granules to settle through a distance of 60 cm in a water column. After the completion of aerobic granulation, the granular sludge was characterized for its partial nitrifying kinetics and settling velocity. For each sludge sample from the bioreactor, the nitritation capability test was performed in a 100 mL glass beaker, and the sludge and NH4+-N concentrations were 2 g MLVSS/L and 400 mg/L, respectively. The sludge mixture was sampled at various time intervals, the NH4+-N, NO2--N and NO3-N concentrations in the liquid phase of the sludge were measured. A first-order correlation could be assumed for the early phase of ammonia degradation in a batch reactor. Therefore, the linear change of ammonia concentration with time per granules biomass can be obtained as the specific rate of ammonia degradation by granules. 2.4. Denaturing gradient gel electrophoresis (DGGE) and bacterial species identification The microbial population was analyzed for the seed sludge and the mature granules. The genomic DNA of the sludge was extracted from the cells using a beadbeater (Mini-beadbeater™, Biospec, Bartlesville, OK, USA) and micro-centrifuge (MiniSpin plus, Eppendorf, Hamburg, Germany) (Zhuang et al., 2005). The bacterial 16S rDNA gene sequence (V3 region, corresponding to positions 341534 of the Escherichia coli sequence) was amplified by polymerase chain reaction (PCR) (PTC-200, MJ Research, Waltham, MA, USA) following the procedure detailed previously (Li et al., 2008). The PCR amplified DNA products were then separated by

DGGE through 8% polyacrylamide gels with a linear gradient of 30-50% denaturant, using the DCode™ Universal Mutation Detection System (Bio-Rad, Hercules, CA, USA). The gels were run for 6 h at 130 V in 1×TAE buffer at 60°C. Afterward, the gels were stained with ethidium bromide for 10 min and then visualized by a UV illuminator. The DGGE images were acquired using the ChemiDoc (Bio-Rad) gel documentation system, and the DGGE profile was analyzed by “QuantityOne” (Version 4.6.3, BioRad, Hercules, CA, USA). A 16S rRNA gene sequence clone library that had been constructed for nitrifying activated sludge was used to identify the phylogeny of the DGGE bands of the sludge samples (Ye and Zhang, 2010). The sequences of the clones used as markers have been deposited in GenBank under accession numbers HM117161 to HM117169. Some of the clones were selected as markers for the DGGE analysis, and the migration positions of the clones were compared with the DGGE profiles of the sludge samples. Based on the comparison, an OTU in the clone was assigned to a particular DGGE band for species identification. 3. Results and discussion 3.1. Formation of aerobic granules for partial nitrification As described previously, daily discharge of slow-settling sludge flocs was applied to the bioreactor. As a result of the selective sludge discharge, aerobic sludge granulation was well achieved fed ammonium as a sole energy source (Fig. 2). It was indicated that the selective discharge of relatively loose sludge flocs is the crucial operating factor for an SBR to achieve granulation (Li et al., 2009). Small and loose sludge flocs were found to have an advantage over larger and dense granules in substrate uptake. The main mechanism of the selective sludge discharge for aerobic granulation is the enhanced feeding of substrates to the biomass of attached-growth. Discharge of loose sludge flocs removes these competitors in suspended-growth mode and makes the substrates more available for uptake by the attached-growth biomass, leading to granulation. The biomass content measured by MLVSS increased gradually from around 2000 to 2300 mg/L (Fig. 3). The mean size of the sludge first increased from 181 to 250 µm after only 12 days, and granule formation became apparent after 2 weeks. The size increased continuously with the formation and growth of granules, and the mature granules had a mean size of around 330 µm (Fig. 4). Based on the lab-scale research results on the cultivation of aerobic granular sludge, it is recommended to develop the aerobic sludge technology further by means of a pilot research at a sewage treatment plant. A high H/D ratio and a certain sequential cycle for SBR may not be necessary for aerobic granulation. Configuration and operation of the

1377


bioreactors can be developed based on the principle of selective discharge of loose and small flocs from system. Moreover, together with the selective sludge discharge method, adjusting the F/M ratio at different stages of granulation can be a more effective start-up strategy (Li et al., 2011). A high F/M ratio can be applied in the early stage, which would bring about fast formation of granules in the reactors. Subsequently, the F/M ratio could be reduced to a lower level to allow the formation and stabilization of smaller granules. Effective control of the size of aerobic granules is of great importance to the development and actual application of the granulation technology in biological wastewater treatment. The concentrations of nitrogen in various forms (e.g. NH4+-N, NO2−-N and NO3−-N) in the effluent from the bioreactor were measured during the experiment. The bioreactor performed well in ammonia removal, as the NH4+-N concentration decreased from around 400 to less than 4 mg N/L (Fig. 5). Initially, complete nitrification was observed in the bioreactor with a low NO2−-N content in the effluent. With the formation of granules, partial nitrification became more predominant and NO3−-N in the effluent dropped to a level below 30 mg N/L. In the early start-up phase, little nitrite accumulated in the reactors, and most of NH4+-N was oxidized to NO3−-N. With the completion of sludge granulation, more nitrite accumulated in the reactors and the nitrate concentration decreased continuously to a low level. After about 30 days, more than 90% of the N content was in the form of NO2−-N and less than 10% of nitrogen was NO3−-N. It is apparent that sludge in the granular form would favor partial nitrification to complete nitrification compared with activated sludge flocs. Sludge granulation can be a method of microbial immobilization used to preserve a high population of nitrifying bacteria within a reactor (Tsuneda et al., 2003). The biomass yields of AOB and NOB were rather low with yield coefficients of 0.14 and 0.072 g/g N, respectively (Blackburne et al., 2007). The selective sludge discharge not only contributed to the formation of granules but also led to more accumulation of AOB than NOB in the granular sludge.

The high-strength ammonia influent would promote more AOB growth. In addition, the larger size of granules would help maintain a low DO condition within the granules, which is unfavorable to the function and growth of NOB. Therefore, sludge granulation provides an environment that is highly favorable to a stable operation of nitritation. As suggested by others (Ji and Chen, 2010; Zhang et al., 2010), nitrite produced by partial nitrification could be readily treated for denitrification by anammox or similar processes. 3.2. Sludge performance and microbial population dynamics during aerobic granulation The sludge settleability and compressibility were improved significantly with the aerobic granulation in the bioreactor. The average settling velocities of mature granules were 2.15±0.90 mm/s, which were much faster activated sludge flocs with a settling velocity of slower than 1.0±0.12 mm/s. Compared with the seed sludge SVI5 of 80.5±8.0 ml/g, the SVI5 were 30.0±2.1 for the granular sludge. Batch tests were conducted on the mature granules for the nitrogen transformation dynamics during ammonium oxidation in the bioreactor (Fig. 6). The results show the reliable performance of the granular sludge in partial nitrification. NH4+-N with a high initial concentration of 400 N/L was completely oxidized within 10 h. More than 95% of NH4+-N was converted to NO2−-N, and the NO3−-N content was only about 5% or less of NO2−-N in the treated effluent. For the initial NH4+-N concentrations of 400 mg N/L, the mature granules had a specific NH4+-N degradation rate of 0.70 g NH4+-N/g VSS·d. Aerobic nitrifying granules was reported to have a specific NH4+-N degradation rate of 0.70 g NH4+-N/g VSS·d (Fang et al., 2009). However, a lower specific NH4+-N degradation rate of only 0.14 g NH4+-N/g VSS·d was reported for granular sludge with a nitritation function treating the sludge reject water (Lopez-Palau et al., 2011). Well-resolved DGGE bands were obtained for the biomass from the bioreactor (Fig. 7). In a mixture of extracted DNA, less abundant sequences may not be amplified sufficiently to form visible DGGE bands (Eichner et al., 1999; Li et al., 2008).

Fig. 2. Image of seed sludge (a) and mature granules after running 30 days (b) produced in the bioreactor (Bar=0.5 mm)

1378

Aerobic sludge granulation for partial nitrification of ammonia-rich inorganic wastewater

Fig. 3. MLVSS concentration of the SBR

Fig. 4. Particle size distribution of sludge in SBR 500 Concentration (mg/L)

Effluent concentration (mg/L)

400 300 NH4+-N NO2--N

200

NO3--N

100

400 300

NH4+-N NO2--N

200

NO3--N 100 0

0 0

10

20

30

40

50

60

0

2

Fig. 5. The ammonia oxidation performance of SBR during sludge granulation: (a) effluent NH4+-N, (b) effluent NO2--N, (c) effluent NO3--N

The DGGE profiles of the sludge samples indicated the dominant bacterial species in the granules. Sludge granulation had an apparent effect on the species selection and accumulation. DGGE profile showed that Band 1 (Nitrosomonas sp., one of AOB species) occurred in the DGGE profiles of the seed sludge and the granular sludge. This evidenced that aerobic granulation through selective sludge discharge could still keep AOB in the bioreactor. The specie indicated by Band 3 was dominant in seed nitrifying sludge; however it disappeared in mature granules. This specie is closely related to the genera Nitrospira sp., which is one of NOB species to oxidize nitrite to nitrate. Thus, sludge granulation did not result in NOB enrichment in the bioreactor. The result of bacterial DGGE analysis is consistent the performance of nitrification of the bioreactor. During the sludge granulation process, the bioreactor was converted from complete nitrification to nitrication with little nitrate in the effluent (Fig. 5). Partial nitrification was achieved and stable for mature granules as AOB became much dominant than NOB in the bioreactor of granular sludge. 4. Conclusions Aerobic granules were successfully cultivated for partial nitrification through selective discharge of small and slow-settling sludge flocs treating inorganic wastewater with high-strength influent of 400 mg NH4+-N/L. Sludge granulation could be prior ammonia oxidation to the level of partial nitrification to complete nitrification, which is highly favorable to a stable operation of partial nitrification.

4

6

8

10

Time (h)

Time (d)

Fig. 6. Concentrations of NH4+-N, NO2−-N and NO3−-N in the batch ammonia oxidation experiment.

According to DNA-base molecular analysis, aerobic granulation resulted in an enrichment of AOB and a reduction of NOB in the granular sludge.

Fig.7. DGGE profiles of the bacterial communities for seed sludge (SS) and granular sludge (GS) (Left: image; Right: schematic)

Acknowledgements This research was supported by grant 51129803 and 51208038 from the Natural Science Foundation of China, grants HKU714811E from the Research Grants Council (RGC) and SEG_HKU10 from the University Grants Committee (UGC) of the Hong Kong SAR Government and National High-Tech Research and Development Program of China (2009AA06XK1483211). The technical assistance of Mr. Keith C.H. Wong and Mr. Lin Ye are highly appreciated.

References APHA, (2005), Standard methods for the examination of water and wastewater, 21st ed. American Public Health Administration, Washington, DC.

1379


Bartroli A., Perez J., Carrera J., (2010), Applying ratio control in a continuous granular reactor to achieve full nitritation under stable operating conditions, Environmental Science and Technology, 44, 89308935. Bautista-Patacsil L., Rollon A.P., Huelgas A.P., Hu J.Y., (2012), Biodegradation and sorption of 17 alphaethinylestradiol in submerged membrane bioreactor: effect of initial ammonium concentration, Environmental Engineering and Management Journal, 11, 1763-1767. Belmonte M., Vazquez-Padin J.R., Figueroa M., Franco A., Mosquera-Corral A., Campos J.L., Mendez R., (2009), Characteristics of nitrifying granules developed in an air pulsing SBR, Process Biochemistry, 44, 602-606. Blackburne R., Vadivelu V.M., Yuan Z., Keller J., (2007), Determination of growth rate and yield of nitrifying bacteria by measuring carbon dioxide uptake rate, Water Environmental Research, 79, 2437-2445. de Graaff M.S., Zeeman G., Temmink H., van Loosdrecht M.C.M., Buisman C.J.N., (2010), Long term partial nitritation of anaerobically treated black water and the emission of nitrous oxide, Water Research, 44, 21712178. de Kreuk M.K., van Loosdrecht M.C.M., (2006), Formation of aerobic granules with domestic sewage, Journal of Environmental Engineering, 132, 694-697. Eichner C.A., Erb R.W., Timmis K.N., Wagner-Döbler I., (1999), Thermal gradient gel electrophoresis analysis of bioprotection from pollutant shocks in the activated sludge microbial community, Applied Environmental Microbiology, 65, 102-109. Fang F., Ni B.J., Li X.Y., Sheng G.P., Yu H.Q., (2009), Kinetic analysis on the two-step processes of AOB and NOB in aerobic nitrifying granules, Applied Environmental Microbiology, 83, 1159-1169. Furukawa K., Inatomi Y., Qiao S., Quan L., Yamamoto T., Isaka K., Sumino T., (2009), Innovative treatment system for digester liquor using anammox process, Bioresource Technology, 100, 5437-5443. Fux C., Huang D., Monti A., Siegrist H., (2004), Difficulties in maintaining long-term partial nitritation of ammonium-rich sludge digester liquids in a moving-bed biofilm reactor (MBBR), Water Science and Technology, 49, 53-60. Han X.L., Hai R.T., Wang W.X., (2010), effect of COD/N ratio and pH on nitrification in a laboratory-scale constructed wetlands treating septic tank wastewater, Environmental Engineering and Management Journal, 9, 615-621. Ji Z.Y., Chen Y.G., (2010), Using sludge fermentation liquid to improve wastewater short-cut nitrificationdenitrification and denitrifying phosphorus removal via nitrite, Environmental Science and Technology, 44, 8957-8963. Jubany I., Lafuente J., Baeza J.A. and Carrera J. (2009) Total and stable washout of nitrite oxidizing bacteria from a nitrifying continuous activated sludge system using automatic control based on Oxygen Uptake Rate measurements. Water Research, 43, 2761-2772. Kim D.J., Seo D., (2006), Selective enrichment and granulation of ammonia oxidizers in a sequencing batch airlift reactor, Process Biochemistry, 41, 10551062.

1380

Li A.J., Li X.Y., (2009), Selective sludge discharge as the determining factor in SBR aerobic granulation: numerical modeling and experimental verification, Water Research, 43, 3387-3396. Li A.J., Li X.Y., Yu H.Q., (2011), Effect of the food-tomicroorganism (F/M) ratio on the formation and size of aerobic sludge granules, Process Biochemistry, 46, 2269-2276. Li A.J., Yang S.F., Li X.Y., Gu J.D., (2008), Microbial population dynamics during aerobic sludge granulation at different organic loading rates, Water Research, 42, 3552-3560. Liu Y., Tay J.H., (2004), State of the art of biogranulation technology for wastewater treatment, Biotechnology Advances, 22, 533-563. Lopez-Palau S., Pericas A., Dosta J., Mata-Alvarez J., (2011), Partial nitrification of sludge reject water by means of aerobic granulation, Water Science and Technology, 64, 1906-1912. Ostace G.S., Cristea V.M., Agachi P.S., (2012), Evaluation of different control strategies of the waste water treatment plant based on a modified activated sludge model no. 3, Environmental Engineering and Management Journal, 11, 147-164. Qin L., Tay J.H., Liu Y., (2004), Selection pressure is a driving force of aerobic granulation in sequencing batch reactors, Process Biochemistry, 39, 579-584. Shi X.Y., Sheng G.P., Li X.Y., Yu H.Q., (2010), Operation of a sequencing batch reactor for cultivating autotrophic nitrifying granules, Bioresource Technology, 101, 2960-2964. Tsuneda S., Nagano T., Hoshino T., Ejiri Y., Noda N., Hirata A., (2003), Characterization of nitrifying granules produced in an aerobic upflow fluidized bed reactor, Water Research, 37, 4965-4973. Turk O., Mavinic D.S., (1987), Benefits of using selective inhibition to remove nitrogen from highly nitrogenous wastes, Environmental Technology Letters, 8, 419426. Van Hulle S.W.H., Vandeweyer H.J.P., Meesschaert B.D., Vanrolleghem P.A., Dejans P., Dumoulin A., (2010), Engineering aspects and practical application of autotrophic nitrogen removal from nitrogen rich streams, Chemical Engineering Journal, 162, 1-20. Yamamoto T., Takaki K., Koyama T., Furukawa K., (2008), Long-term stability of partial nitritation of swine wastewater digester liquor and its subsequent treatment by Anammox, Bioresource Technology, 99, 6419-6425. Ye L., Zhang T., (2010), Estimation of nitrifier abundances in a partial nitrification reactor treating ammoniumrich saline wastewater using DGGE, T-RFLP and mathematical modeling, Applied Microbiology Biotechnology, 88, 1403–1412. 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, 377383. Zhang L., Yang J.C., Furukawa K., (2010), Stable and high-rate nitrogen removal from reject water by partial nitrification and subsequent anammox, Journal of Bioscience and Bioengineering, 110, 441-448. Zhuang W.Q., Tay J.H., Yi S., Tay S.T.L., (2005), Microbial adaptation to biodegradation of tert-butyl alcohol in a sequencing batch reactor, Journal of Biotechnology, 118, 45-53.


July 2013, Vol.12, No. 7, 1381-1391



ADSORPTION OF HEXACHLOROCYCLOHEXANE BY RAW AND SURFACTANT MODIFIED MEERSCHAUM Shengke Yang1,2, Wenke Wang1,2, Yue Zhao1,2, Chanjuan Gao1,2, Yaqian Zhao2,3 1

Key Laboratory of Subsurface Hydrology and Ecological in Arid Region of Ministry of Education, Xi’an 710054, P.R. China 2 School of Environmental Science and Engineering, Chang’an University, Xi’an 710054, P.R. China 3 UCD Dooge Centre for Water Resources Research, School of Civil, Structural and Environmental Engineering, University College Dublin, Belfield, Dublin 4, Ireland

Abstract The adsorption of hexachlorocyclohexane (HCH) by modified meerschaum was studied. The comparison of the meerschaum modified by anion and cation surfactant was conducted to explore their effect on HCH adsorption. The results showed that the absorbability of HCH by the CTMAB-M (cetyltrimethylammonium bromide-modified) is better than that of SAS-M (sodium dodecyl sulphonate-modified) because the CTMAB-M has fluffier fibers structure. The addition of amount of adsorbents should improve the adsorption. When pH value exceeds 11.0, adsorption exhibits optimal behaviour. Results of isotherms studies reveal that the Freundlich isotherm is the best model in current experimental conditions to describe the adsorption behaviour. The adsorption kinetics of NM (natural meerschaum) and CTMAB-M followed the pseudo-first-order kinetic model while SAS-M followed the pseudo-second-order kinetic model. The calculated thermodynamic parameters indicated that the adsorption of CTMAB-M was spontaneous and endothermal in nature. Key words: adsorption, hexachloro-cyclohexane (HCH), meerschaum Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Persistent organic pollutants (POPs) have been recognized as a kind of natural or synthetic organic pollutant with long-term residual property, biological retention, semi-volatilization and high toxicity (Gavrilescu, 2009; Li, 1999). POPs can either move across long distance in atmosphere or sink into the ground (Piao et al., 2004). Environmentalists throughout the world have paid great attentions to POPs as they have serious effects on human health and the environment (Caliman and Gavrilescu, 2009; Muir et al., 1995; Preda et al., 2012). United Nations Environment Programme (UNEP) ever announced 12 kinds of POPs (Shen, 2005), which are the first batch of their kind to be 

announced. As a typical POP, pesticide hexachlorocyclohexane (HCH) was included in the list (Kishimba et al., 2004; Shi et al., 2007). With a halflife as long as 20-50 years in environment, HCH can retain for a long period of time and accumulate in air, water, soil, sediment, plants, animals and the human bodies (Kishimba et al., 2004; Shi et al., 2007; Yang et al., 2007). In addition, HCH can enter organisms via gastrointestinal tract, respiratory tract and skin absorption. Toxicity of γ-HCH is strongest among all kinds of isomers. After entering an organism, HCH mainly accumulates in central nervous system and adipose tissues to stimulate cerebra and cerebella movement. It then leads to pathological reaction on nervous system, digestive system, respiratory system and circulatory system. Moreover, HCH can affect

Author to whom all correspondence should be addressed: E-mail: [email protected]; Phone: +86-29-82339982; Fax: +86-29-85585485


vegetative nervous system and surrounding nerves through cortex. It also affects oxidative phosphorylation of cells in viscera, thus leading to malnutrition, degeneration and putrescence of viscera. In addition, HCH can induce hepatocyte microsome to oxidize enzyme, affect internal secretion activity and restrain ATP enzyme (Wang, 2006; Chen, 1993; Brabdt, 1998). Although HCH had been gradually prohibited in China by the end of 1980s, great concerns are still existing regarding its pollution, movement, transformation, removal and management (Chen et al., 2000; Hu et al., 2006; Lu et al., 2006;Liu et al., 2001; Yang et al., 2003; Zhang and Dong, 2002). The similar situation was faced in other countries and a large number of studies on transformation rule and management of HCH were conducted accordingly (Covaei et al., 2001; Koci et al., 2007). Since POPs can’t be easily degraded and they will detain in environment for a long time, researchers have tried hard to explore the way to highly and efficiently remove them (Dogan et al., 2009; Lin et al., 2011). In China, Wu (2004) has reported a series of achievements in POPs control and pollution remediation using the clay minerals. Many reports (Deryło-Marczewska and Marczewski, 2002; Dultz et al., 2005; Sánchez-Martín et al., 2008) have provided evidences that migration of POPs affected by some factors such as pH, the structure of adsorbent, the layer of electric charge and cation. Consequently, HCH was reckoned that its adsorption behavior was limited by aforementioned factors. Meerschaum is fibrous magnesium-aluminate mineral, which has cellular inner pore and is a kind of natural nano-mineral material with excellent properties (Li et al., 2010; Xie et al., 2009). Organic modification of meerschaum allows the attached hydrophobic organic groups into meerschaum layers, thus transforming hydrophilic meerschaum into lipophilicity. So raw- and modifiedmeerschaum seem to have potential applicable values. The authors’ group also studied the nonmetal mineral materials and their use in pollutants’ (As, F, Cd and Pb) removal (Yang and Wang, 2000; Yang et al., 2000, 2002; Yang and Fei, 2004). However, in this study, HCH immobilization was investigated using raw- and organic (anion/cation surfactants) modified-meerschaum to compare the adsorption abilities of those three meerschaums and to find the most high-efficiency modification. Focuses are placed on the adsorption ability and capacity of removing condition on HCH via organic modified meerschaum. Effect factors and adsorption kinetics are also studied. 2. Experimental section 2.1. Meerschaum and its modification Original meerschaum mineral was obtained from Henan Neixiang Mineral Bureau (Henan, China). During organic modification, cracked

1382

meerschaum was soaked with 1% SAS (sodium dodecyl sulphonate, w/w) and 1% CTMAB (cetyltrimethylammonium bromide, w/w) solution, respectively, for two hours. Thereafter, the meerschaum was filtered, dried, ground and sieved through 60 mesh sieve for future use. 2.2. HCH and other reagents HCH (chromatographically pure) was purchased from National Institute for the Control of Pharmaceutical and Biological Products (Beijing, China). Nitrogen (99.999%) was purchased from Xi’an Messer Ltd. (Xi’an, China). Cyclohexane, CTMAB and SAS were purchased from Xi’an Chemical Reagent Factory (Xi’an, China). All the reagents used were of analytical reagent grades. Deionized water was used throughout the entire experimental study. 2.3. Experimental procedures Stock solutions for sorption studies were prepared by adding 0.02g the HCH (each of α-HCH, β-HCH, γ-HCH and δ-HCH 0.05g) in 10 L deionized water with 5% methanol. After vibrating for a week, the mixture liquid was filtered and the concentration of filtrate was adjusted as 1.62 mg/L. The stock solution was diluted to the targeting concentration for experimental use. For each time 0.5 g meerschaum and 0.27 mg/L 100 mL HCH solution were mixed in a 200 mL beaker, which was then shaken in a thermostat shaker at 125 rpm. The pH was adjusted to 7.0 and the adsorption temperature was set as 250C, while the equilibrium time was set as 24 h, which was determined via the preliminary experiments. After equilibration, the residual HCH in samples was extracted by cyclohexane solution before the samples were filtered using a 0.45 Millipore membrane, then the gas chromatograph (GC) was employed to measure the concentration of remaining HCH. In order to figure out the adsorption isotherm, six levels of initial HCH concentrations (0.10; 0.20; 0.40; 0.60; 0.81; 1.25; 1.62 mg/L) were used while the other adsorption conditions remained constant. The residual HCH in samples was monitored by the aforementioned method. Adsorptions conditions were investigated. Initial pH values of mixture were adjusted to 1.0, 3.0, 5.0, 7.0, 9.0 and 11.0, respectively, by adding hydrochloric acid or sodium hydroxide solutions. The adsorption temperature was controlled from 150C to 350C by Bed Temperature Incubator. Different weight of meerschaum ranged from 0.0 to 10.0 mg/L (0.0; 0.5; 1.0; 2.0; 5.0; 10mg/L) was used and adsorption time was examined from 1 h to 32 h as well. 2.4. HCH monitoring with

A GC—14C gas chromatograph equipped DB-1 capillary column, ECD detector

Adsorption of hexachloro-cyclohexane by raw- and surfactant modified-meerschaum

(Shimadzu Apparatus Manufacturer) was used to monitor HCH. Nitrogen with the purity of 99.999% was employed as carrier gas. To make the HCH standard for celebration purpose, stock solutions of HCH were prepared by dissolving 0.04g HCH (each of α-HCH, β-HCH, γ-HCH and δ-HCH 0.01g) in 1 L cyclohexane. The solution was then diluted to the targeting concentration of celebration purpose. Non-bypass flow injection was used with temperature of injection port of 230 ℃ while the temperature of detector is 300℃. The peak appearing sequence of the groups is α-HCH, β-HCH and γHCH, followed by δ-HCH. In addition, XL30 scanning electron microscope (Netherlands Philips Corporation) and LP115 pH meter (Mettler Apparatus Corporation) were used in the HCH adsorption experiments. 2.5. Isotherm models In the current study, Henry, Langmuir, Freundlich and Temkin isotherms were, respectively, used to fit the sorption data. The linear expressions of those patterns were listed as follows:

Qe =K d Ce

(1)

Ce C 1 = + e Qe Qmax k L Qmax

is

(2)

where: Ce (mg/L), Qe (mg/g), Qmax (mg/g) and kL (L/mg) are equilibrium concentration of solution, adsorption capacity, the maximum adsorption capacity, and Langmuir isotherm constant (which is related to the free energy of sorption process), respectively. The linear plot of Ce Qe versus Ce was used to calculate Qmax and kL from the slope and intercept of the fitting straight line.

1 ln Qe =ln k F + ln Ce n

3. Results 3.1. Effect of amount of adsorbents on HCH adsorption The effect of amount of adsorbents on HCH adsorption is shown in Fig. 1. Results showed that the amount of adsorbents has a positive effect. With the mass of meerschaum increased, immobilization rate of HCH increases rapidly especially for CTMAB-M. The maximum adsorption rate of CTMAB-M is 71.9% at the dosage was 10mg/L as there is marginal increase of the adsorption rate beyond the dosage of 10 mg/L (data not shown). Relatively, the adsorption rate of natural meerschaum (NM) and SAS-M exhibits gentle increase compared with that of CTMAB-M. Their maximum adsorption rates are 36.3% for NM and 11.61% for SAS-M. The possible reason is that SAS was negative-surfactant modified, which may be hard to combine with meerschaum. The addition of SAS may affect the original structure of sepiolite, resulting in the lower scavenging rate 3.2. Effect of time on HCH adsorption

where, Qe (mg/g) is adsorption capacity; Ce (mg/L) is the equilibrium concentration of solution; Kd partition coefficient.

where k T is Temkin constant.

(3)

The effect of time on HCH adsorption is illustrated in Fig. 2. It shows that the CTMAB-M exhibits excellent adsorption behaviour. Within the first 12 hrs adsorption, all three adsorbents show the same trend of adsorption rate being increased with the increase of time. Thereafter, the adsorption seems to quickly reach the equilibrium status. If considering the equilibrium being reached at 24 hrs, the maximum adsorption rate for CTMABM, NM and SAS-M is 57.8%, 25.0% and 9.2%, respectively. It is clear that the CTMAB-M shows good adsorption behavior among the three adsorbents tested. It also reflects that after modification by SAS, due to the effects of surface electric charge and structure, adsorption rate of SAS-M is reduced compared with NM. In the following tests of the current study, 24 hrs was adopted to be the equilibrium time for further adsorption tests. 3.3. Adsorption isotherms

1/n

where, kF (mg/g) (L/mg) and 1/n are the Freundlich constants which are related to the bonding energy and heterogeneity factor of adsorption curve, respectively. Meanwhile, they maybe reveal energy size about the adsorption process and heterogeneity of adsorbents. The linear plot of ln Qe versus ln Ce was used to calculate the 1/n and kF from the slope and intercept of the fitting straight line.

Qe =

RT RT ln A+ ln Ce kT kT

The fitted curves of HCH adsorption by three meerschsums with Henry, Langmuir, Freundlich and Temkin isotherms are illustrated in Fig. 3, and the parameters and corresponding R2 values were computed and listed in Table 1. Although the four isotherms can mathematically describe the adsorption behaviour (Fig. 3 and R2 values in Table 1), it is clear that CTMAB-M shows a highest adsorption capacity with kinetic parameter values of one magnitude higher than NM and SAS-M.

(4)

1383


100

80 NM S A S -M C T M A B -M

80 Removal ratio(%) Removal rate (%)

Removal %) Removalratio( rate (%)

60

40

20

0

60 40 20 0

0

2

NM SAS-M CTMAB-M

4

6

8

10

0

5

10

15

20

25

30

35

Adsorption time (hour) Adsorption time(h)

AAmount m ount of adsorbents (m g/L) of adsorbents (mg/L)

Fig.1. Effect of amount of adsorbents on HCH adsorption

Fig. 2. Effect of adsorption time on HCH adsorption

a)

b)

c) Fig. 3. Isotherms fitting curves of NM (a), SAS-M (b) and CTMAB-M (c)

3.4. Effect of temperature on HCH adsorption Temperature is of importance to all the reactions in which thermal energy participate. This study took CTMAB-M as an example to investigate the effect of temperature (150C, 250C, 350C) on adsorption process. Langmuir and Freundlich isotherms were employed to fit the adsorption isothermal curve at different temperature.

1384

The results are illustrated in Fig. 4. It shows that the adsorption ability in terms of Qe increased with the temperature rising, indicating that high adsorbing temperature favours adsorption reaction. This infers that the adsorption process is an endothermic reaction. The R2 value of the isotherm fitting indicates that Freundlich is more suitable to describe the adsorption behaviour in the temperature range tested.


Table 1. Parameters and errors analysis of adsorption isotherms Parameter Henry isotherm Kd b R2 SSE SAE HYBRID NPSD（%） Langmuir isotherm Qm KL R2 SSE SAE HYBRID NPSD（%） Freundlich isotherm Kf 1/n R2 SSE SAE HYBRID NPSD（%） Temkin isotherm a b R2 SSE SAE HYBRID NPSD（%）

NM

SAS-M

CTMAB-M

0.0340 0.0340 0.9633 0.000012 0.007076 0.003905 3.276011

0.0210 0.0090 0.9670 0.000009 0.007021 0.008059 7.889541

0.2611 0.0468 0.9355 0.000605 0.056059 0.132781 17.231251

0.0480 72.9961 0.6771 0.000066 0.019028 0.022383 7.925477

0.0199 21.1400 0.6589 0.000028 0.011521 0.026670 15.036315

0.1866 8.2264 0.9318 0.000639 0.059956 0.141510 17.571838

0.0549 0.1419 0.9474 0.000011 0.007660 0.003707 3.254483

0.0238 0.2806 0.8918 0.000009 0.006975 0.008041 8.016645

0.2205 0.4398 0.9846 0.000145 0.022970 0.018484 4.357284

0.0530 0.0050 0.9161 0.000029 0.012781 0.010252 5.4423

0.0214 0.0036 0.8173 0.000015 0.008481 0.013215 10.1593

0.1726 0.0360 0.9348 0.000611 0.060422 0.101307 13.3958

3.5. Effect of pH on HCH adsorption Generally, pH is a very significant factor to affect adsorption. The effect of the pH on the HCH adsorption was examined and the results are illustrated in Fig. 5. The striking feature of Fig. 5 is that the pH seems insensitive to the HCH adsorption especially in pH between 1.0 and 9.0 no matter the adsorbents of NM, SAS-M or CTMAB-M being used. However, in strong alkali solution, pH>11, the rate of HCH immobilization was racketed rapidly. The reason behind this may be complicated, but it could be something related with the strong alkali solution as it was so strong that magnesium in meerschaum detached and formed magnesium hydroxide colloid which help adsorbing the HCH of the solution. 3.6. Microcosmic appearance of meerschaum before and after organic modification The scan electron microscope (SEM) has become an important tool during the past decades to examine the conformation of macromolecules. The SEM observation of the natural meerschaum and modified-meerschaum is shown in Fig. 6.

It can be seen from Fig. 6a that NM consists of fibrous crystalloids that are well ranged and ordered to form the smooth surface. The diameter of fibre is about 1-3µm. SAS-M was also observed with distinct fibrous, of which shape, structure and diameter are obviously changed (Fig. 6b). Fig. 6c shows that the fibre of CTMAB-M becomes fluffier. The diameter of fibre increased to 15-25µm and appeared haphazardly. 4. Discussion 4.1. Adsorption kinetics Based on the adsorption tests in this study, it seems that, at the beginning of adsorption, the process belongs to surface adsorption. The voids of adsorbents was sufficient, the speed of adsorption was fast. Then the speed got slow as the adsorption continues, this is because that HCH in the solution migrated and into the inner of meerschaum and scattered, and the residual space became crowd, the sorption speed turned slowing down in this period. Finally, the adsorption reached equilibrium, indicating that the hole/space of adsorbents was

1385


nearly occupied, and the adopting louses of meerschaums are saturated. In theory, the speed of adsorption was nearly equal to that of desorption. Adsorption kinetics is useful to describe this behaviour. Two kinds of kinetic models were employed to explore the relationship between adsorption time and adsorption rate (Lin et al., 2011), namely pseudo-firstorder kinetic model and pseudo-second-order kinetic model. The linearized form of the pseudo-first-order kinetic model is presented as Eq. (5):

The R2 values of adsorption kinetic curves of NM, SAS-M, CTMAB-M calculated by pseudo-firstorder kinetic model are high to some extent (0.92262, 0.85684 and 0.94905 corresponding to NM, SAS-M and CTMAB-M, respectively) while the R2 values from pseudo-second-order kinetic model were 0.64801, 0.98556 and 0.66124 respectively. By inspecting R2, it is clear that the pseudo-first-order kinetic model is suitable to describe the adsorption kinetics data of NM and CTMAB-M, while the pseudo-second-order kinetic model to describe the adsorption kinetics data of SAS-M.

ln  Qe -Qt  =ln Qe -k1t

4.2. RL and error analysis of adsorption isotherms fitting

(5)

where, Qe (mg/g) is equilibrium adsorption capacity of meerschaums to HCH; Qt (mg/g) is adsorption

A dimensionless constant separation factor RL was used to express the essential characteristics, which is defined as follows:

capacity in t hour; k1 (1/min) is the rate constant of pseudo-first-order kinetic model. The linear plot of ln Qe  Qt  versus t could calculate k1 and Qe , cal

RL =

from slope and intercept, respectively. The linearized form of the pseudo-secondorder kinetic model is presented as Eq. (6):

where C0 (mg/L) is the initial concentration of HCH

t 1 t = + 2 Qt k2Qe Qe

(6) where, k2 (1/min) is the rate constant of pseudosecond-order kinetic model. The linear plot of t Qt versus t was employed to calculate k1 and

Qe, cal from

intercept and slope, respectively. Both pseudo-first-order kinetic model and pseudosecond-order kinetic model were used to fit the kinetic data of meerschaums adsorbing HCH and the fitting charts are shown in Fig. 7, and the parameters of corresponding equations are listed in Table 2.

a)

1 1+k LC0

(7)

in solution. The adsorption is considered as irreversible when RL =0, favorable when 0 < R < 1, linear when RL = 1, and unfavourable when RL >1(Liu et al., 2010). In this study, the RL values for NM, SAS-M, CTMAB-M were 0.048 ， 0.149 and 0.310, respectively, which were all between 0 and 1, suggesting that the adsorptions were favourable. Generally speaking, the R2 is widely used to examine whether the model are accord with the experimental data, and the value is nearer to unity being deemed to provide the best fit. However, due to the inherent bias resulting from linearization, R2 is not the only criterion to evaluate the best fit (Kundu and Gupta, 2006; Zhao et al., 2007).

b)

Fig. 4. Fitting curves of Langmuir (a) and Freundlich (b) isotherms at different temperature

1386


100 90

NM S AS -M C T M A B -M

Removal ratio( %) Removal rate (%)

80 70 60 50 40 30 20 10 0

0

2

4

6

8

10

12

pH Fig. 5. Effect of pH on HCH adsorption

a)

b)

c)

Fig. 6. SEM observation of HCH before and after modification: a) NM; b) SAS-M; c) CTMAB_M

a.

-3

b. 7000

-4

6000

-5 -6

5000

-7

4000

-8

t/Qe

ln(Qe-Qt)

NM SAS-M CTMAB-M Fit of SAS-M Fit of NM Fit of CTMAB-M

NM SAS-M CTMAB-M Fit of SAS-M Fit of NM Fit of CTMAB-M

-9 -10 -11 0

5

10

3000 2000 1000

15

20

25

30

35

0

0

5

10

t ( h)

15

20

25

30

35

t ( h)

Fig. 7. Pseudo-first-order and pseudo-second-order curves

1387


Table 2. Kinetic functions and fitting parameters First-order kinetics Secondorder kinetics

YNM=-3.48752-0.20389 X YSAS=-5.1885-0.1537 X YCTMAB=-3.25485-0.13963 X YNM=1114.32202+33.63654 X YSAS=1226.20284+158.92288 X YCTMAB=449.47515+13.82945 X

In order to analysis fitting degree of three isotherms for the experimental data, some errors functions are often used to integrated evaluate the fit (Zhao et al., 2007). The sum of the squares of the errors (SSE): n

2

i=1

i

SSE    Qe ,cal -Qe ,exp 

(8)

The sum of the absolute errors (SAE):

i=1

The hybrid (HYBRID):

i

(9)

fractional

error

function

2   100 n   Qe,cal -Qe,exp   HYBRID    n-p i=1  Qe,exp   i (10)

Normalized percentage standard deviation (NPSD)：

NPSD(%)  100

1 n  Qe,cal -Qe,exp  n-1 i=1  Qe,exp

where, Qe , exp and Qe , cal

2

  i (11)

are respectively the

experimental values and calculated values by the adsorption isotherm; n and p refer to the number of data points and the number of isotherm parameter, respectively. Isotherm parameters with error analysis and the magnitude of the correlation coefficient R2 are jointly summarized in Table 1. By carefully inspecting Table 1, it is clearly known that the R2 value of Henry is highest to NM (R2=0.9633) and SAS-M (R2=0.9670), suggesting that the first order, Henry model is well suitable to NM and SAS-M. While to CTMAB-M, the R2 of Freundlich is nearest to unity (R2=0.9846), proving that sorption process of CTMAB-M was in line with Freundlich isotherm. The four errors of Henry fitting for NM as well as SAS-M are lowest while the errors of Freundlich fitting for NM and SAS-M are also lower. Hence,

1388

Qe.cal(mg/g) 0.030577 0.00558 0.038587 0.02973 0.006292 0.072309

R2 0.92262 0.85684 0.94905 0.64801 0.98556 0.66124

Qe.exp(mg/g) 0.0135 0.005 0.032

Henry model was much better to describe the sorption process of NM and SAS-M. But to CTMAB-M, the four errors are minimum, which agreed with the results of R2 analysis. According to the four errors and R2 of Langmuir, the results of Langmuir didn’t well match with the experimental data, indicating that the adsorption can’t simply explained by monolayer adsorption theory, especially the adsorption of CTMAB-M, which could be better explained by multilayer adsorption theory instead. 4.3. Adsorption thermodynamics

n

SAE   Qe ,cal -Qe ,exp

k 0.20389 0.1537 0.13963 0.030186 0.129606 0.030768

The parameters of Langmuir and Freundlich isotherms at different temperature are listed in Table 3. As the Freundlich isotherm appeared to be the best fitting, CTMAB-M adsorption and Freundlich constant were used to study the thermodynamics parameters such as Gibbs free energy change (ΔG), enthalpy change (ΔH) and entropy change (ΔS) (Huang et al., 2009; Bell and Tsezos, 1987; Zhang et al., 2010).

G=-nRT lnk L =S / R   H / RT

G   H  T S

(12) (13) (14)

where, R (8.314 J/mol·K) is the gas constant; T (K) is the absolute temperature; the n is Freundlich constant. The values of H and H rise and fall gently, while the values of ΔG varied largely, thus, the values of H and S can be calculated from the intercept and the slope of the linear plot of the straight line of Eq. (14) in ΔG versus T. The obtained values of thermodynamic parameters for CTMAB-M were listed in Table 4. As shown in Table 4, the values of ΔG were always negative at all the investigating temperature. The higher the adsorption temperature, the bigger absolute value of ΔG. It betrayed the feasibility of the adsorption of CTMAB-M and the spontaneous nature of the sorption process. The higher of the adsorption temperature, the better the spontaneous nature. The adsorption was considered to be physisorption dominating when the value of ΔG is between -20 and 0 kJ/moL, and to be chemisorption dominating when the value of ΔG is between -400 and 80 kJ/moL (Yu et al., 2004). In this study, the values of ΔG were less than -5 kJ/moL, indicating that physisorption dominated the adsorption process.


Table 3. Parameters of two adsorption isotherms at different temperatures Parameter Langmuir isotherm Qm KL R2 Freundlich isotherm Kf 1/n R2

288K

298K

308K

0.1844 6.8966 0.9395

0.1869 8.1416 0.9324

0.1965 8.9392 0.9333

0.2063 0.4525 0.9705

0.2204 0.4410 0.9854

0.2379 0.4342 0.9873

Table 4. Parameters of adsorption thermodynamics

 H (kJ/moL)

 S (J/moL)

3.4857

30.5

The value of S obtained in this study is positive (30.5 J/mol·K), showing an increase in the adsorption process of CTMAB-M to adsorb HCH. That’s because the increase of entropy resulting from desorption of molecule of solvent is larger than the decrease of entropy derived from the adsorption of molecule of solvend (Zhou and Zhang, 2010). The energy of adsorption from different forces were as follows: dipole bond forces 2–29 kJ/mol, hydrogen bond forces 2–40 kJ/mol, Van der Waals forces 4–10 kJ/mol, hydrophobic bond forces about 5 kJ/mol, coordination exchange about 40 kJ/mol and chemical bond forces > 60 kJ/mol (Huang, 2008). The value of ΔH was 3.4857 kJ/mol, indicating that hydrogen bonding and dipole bond forces are significant to the adsorption process in addition to electrostatic attraction. The positive value of H indicated that the adsorption process was endothermal in nature, and high temperature was in favour of adsorption. Heating was beneficial of HCH adsorbing to surface of microballoon, promoting the adsorption progress, which conformed to the results of fitting curve. To some extent, it can be declared that CTMAB-M adsorbing HCH was pysisorption and chemisorption process, i.e. the joint actions. Elevating temperature can active adsorption site, but also decrease the activation energy of sorption, which is favorable of forming chemical bond, thus to promote the chemisorption (Zhou and Zhang, 2010). 4.4. The role of meerschaum modification Meerschaum is magnesium-aluminate nature mineral with specific surface area and fibrous appearance. Meerschaum assumes to have electro negativity due to composition of silica skeleton. For this reason, it can adsorb cation well. With negative charge, meerschaum expresses repulsive force to SAS when meerschaum was activated by anionic surfactants. Hence, SAS is hard to stick to the surface of meerschaum.

G (kJ/moL) 288K -5.29

298K -5.62

308K -5.9

Consequently, the surface, structure and diameter of fibre of SAS modified-meerschaum appeared little alteration. This coincides with the result of electron microscope (Fig. 6b). It was reasonable to activate meerschaum with CTMAB due to CTMAB with positive charge. Meerschaum could be reorganized with CTMAB loaded, and display stronger adsorbility to organic matter (HCH). Compared with the raw meerschaum, the meerschaum modified with CTMAB has fluffier fibres, and the diameter of typical fibre increases to 15-25µm. It is noted from Fig. 6c that there are more exiguous irregular fibres on the surface of the CTMAB modified meerschaum. This structure makes meerschaum looser with relatively bigger apertures, thus having considerably better adsorption ability. It may be because that CTMAB modifier entered into interlayer of crystalloid of meerschaum, and the interlayer was unfurled, expanding the spacing between layers of crystal. This can explain that the structures of CTMAB-modified meerschaum become loosen. The looser the structures become, the bigger porosity gets, the larger the specific area represents, the stronger adsorption capacity reflects compared with NM. Zhou and Zhang (2010) has studied the HDTMA.Br-modified vermiculite, believed that CTMA.Br inserted into the multilayer, resulting in the particle exchange and specific area enlargement. In the current study, the features of CTMAB are similar with HDTMA in the study of vermiculite, so the effect could be the similar with HDTMA.Brmodified vermiculite. 4.5. Possible application Although the adsorption mechanisms of HCH onto meerschaum may be complicated and this study cannot fully explore and discuss it, it is clear that the adsorption capacity largely depends on meerschaum’s structure as there are large quantities of adsorption sites in the micro voids. Modification

1389


has been demonstrated as a value-added approach to enhance the adsorption capacity. Therefore, it is believe that CTMAB-modified meerschaum has excellent adsorption performance and can be used to repair HCH-affected environment including soil, surface water and underground water by imbedding into the soil and, alternatively, filling the permeable reactive barrier (PRB) to immobilize HCH in different ways. 5. Conclusions In the present study, the effects of amount of adsorbent, adsorption time, temperature and pH were investigated with kinetics, thermodynamics and adsorption isotherms being studied. The following conclusions can be drawn: 1) Modification of meerschaums is a wise way to promote its HCH adsorption capacity. The order of adsorption ability of meerschaums from high to low is CTMAB-M > NM > SAS-M. 2) The value of H was positive, indicating that adsorption process was endothermic in nature, and high temperature was in favor of adsorption. The values of ΔG were always negative, suggesting that adsorption was feasible and spontaneous nature. 3) The adsorption processes of NM and CTMAB-M can be described by pseudo-first-order kinetic model while the adsorption behaviour of SAS-M can be better described by pseudo-secondorder kinetic model. 4) The study has the potential application in large scale for HCH control. The application could be in reparation of HCH-affected environment including soil, surface water and underground water by imbedding into the soil and, alternatively, filling the permeable reactive barrier. Acknowledgements Funding for this study were provided by the National Natural Science Foundation of China (No. 41072185 and 41372259) and the Special Fund for Basic Scientific Research of Central Colleges (Chang’an University, 2013G3292016 and 2013G1502037).

References Bell J.P., Tsezos M., (1987), Removal of hazardous organic pollutants by biomass adsorption, Journal of Water Pollution Control Federation, 59, 191-198. Brandt I., (1998), Developmental and reproductive toxicity of persistent environmental pollutants, Archives of Toxicology, 20, 111 – 119. Caliman F.A., Gavrilescu M., (2009), Pharmaceuticals, personal care products and endocrine disrupting agents in the environment – a review, CLEAN – Soil, Air, Water, 37, 277–303. Chen R., (1993), The Experiment of Environmental Monitoring, Shanghai Tongji University Press, Shanghai. Chen W., Hong H., Zhang L., Xu L., Wang X., Hong L., (2000), Residue levels and distribution feartures of organochlorine pollutants in the surface sediments of

1390

the sea area between Minjiang Estuary and Mazu, Marine Science Bulletin, 19, 53-59. Covaei A., Hura C., Sehepens P., (2001), Selected persistent organochlorine polutants in Romania, The Science of the Total Environment, 280, 143 -152. Deryło-Marczewska A., Marczewski A.W., (2002), Effect of adsorbate structure on adsorption from solutions, Applied Surface Science, 196, 264-272. Dogan M., Abak H., Alkan M., (2009), Adsorption of methylene blue onto hazelnut shell: Kinetics, mechanism and avtivation parameters, Journal of Hazardous Materials, 164, 172-181. Dultz S., Riebe B., Bunnenberg C., (2005), Temperature effects on iodine adsorption on organo-clay minerals: II. Structural effects, Applied Clay Science, 28, 17-30. Gavrilescu M., (2009), Behaviour of persistent pollutants and risks associated with their presence in the environment - integrated studies, Environmental Engineering and Management Journal, 8, 1517-1531. Hu J., Dou Y., Zhao Z., Zhang J., Liu J., Zhao L., (2006), The environmental impact analysis on phasing out DDT in China, Environmental Pollution & Control, 28, 222 -225. Huang J.H., Liu Y.F., Wang X.G., (2008), Selective adsorption of tannin from flavonoids by organically modified attapulgite clay, Journal of Hazardous Materials, 160, 382-387. Huang J., Huang K., Yan C., (2009), Application of an easily water-compatible hypercrosslinked polymeric adsorbent for efficient removal of catechol and resorcinol in aqueous solution, Journal of Hazardous Materials, 167, 69-74. Kishimba M.A., Henry L., Mwevura H., Mmochi A.J., Mihale M., Hellar H., (2004), The status of pesticide pollution in Tanzania, Talanta, 64, 48-53. Koci V., Ocelka T., Dragoun D., (2007), Concentration of organochlorine pollutants in surface waters of the central European biosphere reserve Krivoklatsko, Environmental Science and Pollution Research, 14, 94-101. Kundu S., Gupta A.K., (2006), Arsenic adsorption onto iron oxide-coated cement (IOCC): Regression analysis of equilibrium data with several isotherm models and their optimization, Chemical Engineering Journal, 122, 93-106. Li J., Li Y., Ma Y., Wang H., Cao W., Wang C., (2010), Adsorption of organic sepiolite to methyl orange from aqueous solutions, Non-Metallic Mines, 33, 67-70. Li Z., (1999), Trends and strategies of global control over persistent organic pollutants, Modern Chemical Industry, 19, 5-8. Lin J., Zhan Y., Zhu Z., Xing Y., (2011), Adsorption of tannic acid from aqueous solution onto surfactantmodified zeolite, Journal of Hazardous Materials, 193, 102–111. Liu G., Jiang G., Zhang H., Lin X., (2001), Long-term dynamics of organochlorined pesticides in unsaturated zone and groundwater, Agro-environmental Protection, 20, 408-417. Liu Q.S., Zheng T., Wang P., Jiang J.P., Li N., (2010), Adsorption isotherm, kinetic and mechanism studies of some substituted phenols on activated carbon fibers, Chemical Engineering Journal, 157, 348-356. Lu A., Yang R., Jiang G., Yang X., Xiong J., (2006), The distributions of organmchlorine pesticides insediments and sewage irrigated area soils collected from Urumqi Shuimohe Valley, Environmental Chemistry, 25, 508510.


Muir D.C.G., Grift N.P., Lockhart W.L., Wilkinson P., Billeck B.N., Brunskill G.J., (1995), Spatial trends and historical profiles of organochlorine pesticides in Arctic Lake sediments, The Science of the Total Environment, 160, 447-457. Piao X., Wang X., Tao S., Shen W., Qin B., Sun R., (2004), Vertical distribution of organochlorine pesticides in farming soils in Tianjin Area, Research of Environmental Sciences, 17, 26-29. Preda C., Ungureanu M.C., Vulpoi C., (2012), Endocrine disruptors in the environment and their impact on human health, Environmental Engineering and Management Journal, 11, 1697-1706. Qu Y., Zhang C., Zhou Q., (2008), Organic pollutant removal from wastewater by surfactant-modified zeolite, Industrial Water Treatment, 28, 17-19. Sánchez-Martín M.J., Dorado M.G., del Hoyo C., Rodríguez-Cruz M.S., (2008), Influence of clay mineral structure and surfactant nature on the adsorption capacity of surfactants by clays, Journal of Hazardous Materials, 150, 115-113. Shen P., (2005), Stockholm Convention and persistent organic pollutants (POPs), Chemical Education, 6, 610. Shi S., Zhou L., Shao D., Huang Y., (2007), Studies on residues of organochloride pesticides POPs in the soils in Beijing area, Research of Environmental Sciences, 20, 24-29. Wang L., (2006), Advances in Chemistry of Organic Pollutants, Chemical Industry Press, Beijing. Wu P., (2004), Clay Minerals and Environmental Modification, Beijiang Chemical Industry Press, Beijing. Xie Z., Chen Z., Dai Y., (2009), Preparation of sepiolite complex sorbent and its treatment properties for dyeing wastewater, Environmental Science &Technology, 32, 130-133. Yang G., Wan K., Zhang T., Guo Z., Wan H,, Luo W,, Gao Y., (2007), Residues and distribution characteristics of organochlorine pesticides in agricultural soils from

typical areas of Guangdong Province, Journal of Agro-Environment Science, 26, 1619-1623. Yang S., Fei X., (2004), Study on effect factors of treating cadmium with meerschaum, Industrial Minerals and Processing, 33, 16-19. Yang S., Wang W., (2000), Study on action and mechanism of removing fluoride with activated meerschaum, Industrial Minerals and Processing, 29, 1-4. Yang S., Wang W., Li X., (2000), Treatment of water arsenic bearing by modified meerschaum, Industrial Minerals and Processing, 29, 13-17. Yang S., Zhou C., Zhang W., et al. (2002), Study on effect factor of treating lead with meerschaum, Industrial Minerals and Processing, 31, 11-15. Yang Y., Liu M., Xu S., Hou L., Liu H., (2003), Distribution of polychlorinated biphenyls (PCBs) and organochlorine pesticides (OCPs) in the tidal beach surface sediments of Yangtze estuary, China Environmental Science, 23, 215-219. Yu Y., Zhuang Y.Y., Wang Z.H., Qiu M.Q., (2004), Adsorption of water-soluble dyes onto modified resin, Chemosphere, 54, 425-430. Zhang J., Huang Y., Liu W., Liang H., (2010), Studies on the adsorption of aromatic compounds onto adsorbent resin, Journal of University of Science and Technology of China, 40, 706-711. Zhang X., Dong X., (2002), Organic chlorinated pesticides in middle and lower reaches of Liaohe, Journal of Dalian Institute of Light Industry, 21, 102-104. Zhao Y.Q., Razali M., Babatunde A.O., Yang Y., Bruen M., (2007), Reuse of aluminum-based water treatment sludge to immobilize a wide range of phosphorus contamination: Equilibrium study with different isotherm models, Separation Science and Technology, 42, 2705-2721. Zhou L., Zhang Z., (2010), Preparation of anion-cation organo verm iculite and study on adsoropion thermodynamics to reactive brilliant red, Journal of Shenyang Ligong University, 29, 79-84.

1391

The Romanian Academy “Petru Poni” Institute of Macromolecular Chemistry, Iaşi 41A, Grigore Ghica Vodă Alley, 700487 Iaşi Tel.: 0232-217454, *260332, *260333, *260334 Fax: 0232-211299; E-mail: [email protected]

Petru Poni Institute of Macromolecular Chemistry, Iaşi – an institute of the Romanian Academy – is the beneficiary of a project financed by structural funds (ID 88, Nr. 03/01.03.2009, Cod SMIS – CSNR 2213), within the framework of the Sectoral Operational Programme “Increase of Economic Competitiveness”, Priority Axis 2 – Operation 2.2.1 “Developing existing R&D infrastructure and creating new R&D infrastructure”. The objective of the project, whose non-refundable value amounts to 15 658 593 lei (equivalent to about 3.7 million Euro), is to construct and provide with high-performance equipment a Centre of Advanced Research in Nanobioconjugates and Biopolymers (IntelCentre), integrated into “Petru Poni” Institute of Macromolecular Chemistry. The research activity of the centre will be focused on: -

controlled release drugs, non-viral vectors, sensors for DNA quatification in solution, nanosupports for fast peptide synthesis.

IntelCentre, having a useful area of about 1000 m2, will comprise 12 laboratories of chemistry, biochemistry, biology, instrumental analysis and molecular characterization. Over 100 last-generation research equipments, especially designed for the synthesis and structural characterization of nanosystems with bioactive components, will be purchased and will serve the future centre. In addition to the advanced research activity in an emerging domain of nanotechnology, IntelCentre will exert an important educational role, offering academic interdisciplinary specialization opportunities to chemists, biologists and physicists, involved in doctoral and postdoctoral programs.


July 2013, Vol.12, No. 7, 1393-1399



ADSORPTION OF PHOSPHATE FROM AQUEOUS SOLUTIONS BY THERMALLY MODIFIED PALYGORSKITE Jingjing Xie1, Tianhuhu Chen1, Chengsong Qing1, Dong Chen1, Chengzhu Zhu1, Jiayuan Wang1, Xinmin Zhan1,2 1

Hefei University of Technology, School of Resources and Environmental Engineering, Hefei, P.R. China 2 National University of Ireland, Department of Civil Engineering, Galway, Ireland

Abstract In this study, palygorskite clay calcinated at different temperatures from 200 to 900 °C was used as adsorbent to remove phosphate from aqueous solutions. The crystal structure and the surface property of the thermally modified material was investigated to determinate the relationship between the adsorption capacity of the adsorbent and the changes of the crystal structure and surface property. The results indicate that palygorskite clay calcinated at 700 °C had the highest adsorption capacity. Complete dehydroxylation destroyed the crystal structure and Al coordination was changed from octahedral to tetrahedral and further to a small amount of penta-coordination. The thermally modified adsorbent material was amorphous. All these structure changes enhanced the activity of Al and phosphate adsorption. Key words: adsorption, palygorskite, phosphate, thermal treatment Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Phosphorous is a nutrient in surface water due to over use of fertilizers and detergents, which results in the over growth of aquatic plants and algae, as well as depletion of dissolved oxygen in the water (Bowden et al., 2009; Shin and Han, 2004). Studies have shown that several main inland lakes in China, such as Tai Lake, Chao Lake and Dianchi Lake, are suffering from serious eutrophication, and the water environmental capacity is exhausted (Chen et al., 2003; Huang et al., 2005; Gao et al., 2005). Phosphorus contained in industrial and domestic wastewaters is also responsible for eutrophication (Peleka and Deliyanni, 2009). Municipal wastewater may contain phosphate from 4 to 15 mg/L, and industrial wastewaters (like in detergent manufacturing wastewater and metal coating wastewater) may contain a phosphate level in excess of 10 mg/L (Akay et al., 1998). 

Phosphate removal from aqueous solutions has been studied for the past decades. Conventional phosphate removal technologies include chemical treatment, biological treatment, and a combination of chemical and biological treatment (Ayoub et al., 2001; Babatunde et al., 2010; Blaney et al., 2007; Donnert et al., 1999; Gustafsson et al., 2008; Jenkins et al., 1971; Karageorgiou et al., 2007; Lee et al., 2009; Zelmanov and Semiat, 2011). In chemical treatment, a divalent or trivalent metal salt is added to wastewater, causing precipitation of an insoluble metal phosphate that is settled out by sedimentation. The commonly used salts are iron and aluminium in the presence of chlorides or sulphates (Hsu, 1975; Fytianos et al., 1998; Zhou et al., 2008; Seida et al., 2002). Alternatively, lime is also used due to the formation of calcium phosphate (Hosni et al., 2008). However, those methods generally produce a large amount of excess sludge with a high water content, which is

Author to whom all correspondence should be addressed: E-mail: [email protected],; Phone: +86 5512903990

Xie et al./Environmental Engineering and Management Journal 12 (2013), 7, 1393-1399

hard to dispose of and will cause secondary contamination (Morse et al., 1998). Biological treatment can avoid the use of chemicals and excess sludge production, and 90% phosphorus can be removed under suitable conditions. However, biological treatment has defects such as poor stability, strict operation, ease influence by temperature and pH, strong dependence on BOD, and so on. The effluent discharge standard can hardly be imposed when the phosphorus concentration is in excess of 10 mg/L. Therefore, the effective phosphate removal from effluent after biological treatment should be studied (Wu et al., 2011). Another problem associated with these methods is that the recycle of phosphorus is not achievable. Therefore, adsorption treatment has attracted more attention (Hano et al., 1997; Donnert et al., 1999; Oguz, 2004; Xiong et al., 2008; Bia et al., 2012; Zhang et al., 2012; Jutidamrongphan et al., 2012). The application of low-cost adsorbents in wastewater treatment has been widely investigated during recent years, such as aluminum hydroxide (Tanada et al., 2003; Genz et al., 2004), fly ash (Agyei et al., 2002), blast furnace slag (Johansson et al., 2000), red mud (Zhao et al., 2012), alunite (Özacar, 2003; Yang et al., 2006), iron oxide tailings (Zeng et al.,2004; Xie et al., 2007) and silicate minerals (Hrenović et al., 2012; Zamparas et al., 2012 ). Palygorskite is a natural magenesiumaluminum silicate clay mineral with a diameter of 30 - 50nm (Chen et al., 2011). It is known to contain continuous two-dimensional tetrahedral sheets, but differs from other layer silicates due to its lack of continuous octahedral sheets. The tetrahedral basal oxygen atoms invert apical directions at regular intervals coordinating talc like ribbons. It has high surface area, high viscosity, high porosity and significant thermal resistance as well as chemical inertness. Palygorskite resource is abundant in eastern Anhui Province and western Jiangsu Province, China (Chen et al., 2004). It is thought as an excellent absorbent and more attention has been paid to the utilization of palygorskite (Murray, 2000; Chen et al., 2000; Peng et al., 2006; Álvarez-Ayuso and García-Sánchez, 2007; Al-Futaisi, 2007). Some researches show that chemical and thermal treatment of palygorskite can increase its capacity in phosphate adsorption (Chen et al., 2010; Ye et al., 2006; Gan et al., 2009). Gan et al. (2009) found that high concentration phosphate might be removed by precipitation of Ca phosphate compounds, and attributed the adsorption capacity increase of palygorskite clay calcined to the changes in the crystal structure of palygorskite and the activation of calcium, iron and aluminium. In comparison with phosphorus removal at high phosphate concentrations through chemical precipitation, the adsorption mechanism of phosphate at low P concentrations by palygorksite clay calcinated is unclear.

1394

The aim of the present work was to study the phosphate removal efficiency of palygorskite thermally modified under different temperatures, and to explore the relationship between adsorption property and crystal structure of palygorskite adsorbents. 2. Materials and methods 2.1. Materials Raw palygorskite clay used in the experiments was obtained from Guanshan Palygorskite Clay Mine, Mingguang, Anhui Province, China. The sample was characterized by XRD and TEM. The XRD analysis of the sample (Fig. 1) revealed the mineral phase of palygorskite accompanied with a small amount of impurities like dolomite and quartz. The chemical composition in weight percentage of oxides (%) is: SiO2 (60.15); MgO (10.56); Al2O3 (9.08); Fe2O3 (4.63); P2O5 (1.31); CaO (2.39); TiO2 (0.65); K2O (0.77); Na2O (0.04); and loss on ignition (10). The image obtained from the transmission electron microscopy (TEM) (Fig. 2) revealed abundant fibers in the palygorskite clay sample. Individual palygorskite fiber was 30 - 40 nm in diameter, and 0.3 - 2 μm long. Before thermal treatment, the palygorskite samples were treated with HCl (2mol/L) under different solid-liquid ratios (1:1, 1:2, 1:3, 1:4), with constant stirring for 20 min by a high-speed blender at 1000 r/min. The sample was then washed to Clfree with distilled water and vacuum dried at 80°C for 24 to 48 h. Then it was manually crushed and screened into particles with specified diameters. Then, the pre-treated palygorskite samples were calcinated in a tubular furnace at temperatures of 200 °C, 400 °C, 450 °C, 500 °C, 550 °C, 600 °C, 700 °C, 800 °C and 900 °C for 2 h, respectively. Thermally modified palygorskite samples were collected. Phosphate solutions were prepared by dissolving potassium dihydrogen orthophosphate (KH2PO4, analytical reagent grade) in distilled water. Then, the pH value of the PO43- solutions was adjusted to 6.9-7.1 with diluted HCl and NaOH. 2.2. Characterization and analysis The structure and textural characteristics of the samples were investigated by XRD, MAS-NMR and BET-N2. XRD patterns of the prepared samples were measured on a Rigaku D/Max-rB X-ray diffractometer (Rigaku, Japan) using Cu Kα radiation (40kV, 100mA). All XRD patterns were obtained from 3.0º to 70º with a scan speed of 6º/min. TEM images were taken by a JEOL 2010 TEM (Jeol, Japan). The NMR spectra of the samples were recorded on a Bruker AVANCE III 400 Nuclear Magnetic Resonance Spectrometer (Buchi, Switzerland) equipped with a magic angle sample spinning rotor at 104MHz.

Adsorption of phosphate from aqueous solutions by thermally modified palygorskite

Fig. 1. XRD pattern of Mingguang palygorskite clay

Fig. 2. TEM image of palygorskite

27

Al chemical shifts were recorded with respect to the solution of AlCl3 as an external reference. The BET surface area was measured by N2 adsorption-desorption technique using a Quantachrome NOVA 3000e analyzer (Quantachrome Instruments, USA). Phosphate concentrations in water samples were determined via ammonium molybdate spectrophotometric method (Moulder et al., 2004). Each analysis point was an average of three independent parallel measurements. Triplicate tests showed that the standard deviation of the results was ±5%. The phosphorous elemental speciation was characterized by X-ray photoelectron spectroscope (XPS). An ESCALAB 250 XPS machine (ThermoVG Scientific，USA) was used for the analysis. The instrument setting was as follow: monochromatic Al Kα radiation (hv = 1486.7 eV) operated at 150W, beam spot of 500μm, and analyzer energy of 20 eV. 2.3. Static adsorption experiment Adsorption measurements were evaluated using batch experiments at ambient temperatures (25±1°C). 0.1 mg of thermally modified palygorskite clay samples were added into 250 mL flasks, and 5 mg P/L solutions were then added to the flasks with 50 mL in each flask. The initial pH of the solution was maintained at 6.9-7.1 by manually adding HCl and NaOH solutions. The flasks were capped and stirred in thermostat water bath at a speed of 180 r/min for 24 h to ensure approximate adsorption equilibrium. At the end of the experiment, the solutions were filtered through 0.45 μm membrane filter paper and then analyzed for PO43concentrations. The quantity of adsorbed phosphate was calculated from the decrease of the phosphate concentration in solutions. The duplicate experiments demonstrated the high repeatability of this adsorption experiment and the experimental error was controlled within 5-10%.

2.4. Column adsorption experiment In column flow-through adsorption, 30 g adsorbent with a particle size of 0.6 - 0.9 mm diameter prepared under the best condition was filled in a dynamic reaction column with an inner diameter of 10 mm and filling height of 580 mm. Phosphate working solution had with a phosphate concentration of 1 mg P/L and was stored in a 10-L PVC plastic bucket. Phosphate solution was fed into the reaction column by a constant flow wriggle pump at a flow rate of 0.46 m/h. The hydraulic residence time was 1.5 h. The phosphate concentration in effluent and the accumulative effluent volume were monitored regularly until the effluent phosphrous concentration was over 0.025 mg/L, indicating the breakthrough of the adsorption column. The sample in the reaction column was collected, and then dried for characterization. 3. Results and discussion 3.1. Effects of calcinated temperature on phosphate adsorption Fig. 3 shows that the adsorption efficiency of palygorskite clay thermally modified after acid treatment was improved obviously in comparison with raw palygorskite clay. Among them the acid treated sample at the solid-liquid ratio 1:3 (1 g palygorskite clay: 3 mL HCl) had the highest adsorption efficiency. The calcinations temperature of the acid treatment samples for the highest adsorption effect is 100°C lower than those samples without acid treatment. The adsorption capacity of palygorskite thermally modified at different temperatures varied from each other. When the calcination temperature was lower than 400 °C, the thermal treatment had no significant influence on adsorption; while the adsorption capacity was improved with the rise of 1395


calcination temperature when the calcinations temperature was between 400 °C to 700 °C. The adsorption capacity reached the highest when the calcination temperature was between 600 °C to 700 °C, and the temperature window of thermal activation was the broadest at this range. Therefore the optimal modification conditionwould be as follows: the solid-liquid ratio for acid pretreatment is palygorskite: HCl of 1:3, and calcination temperature is 700 °C. Therefore, the adsorbent used in the following adsorption experiment was manufactured under this optimal condition. 3.2. Changes of crystal structure and specific surface area Fig. 4 shows the XRD patterns of raw and modified palygorskite clay. Compared with raw palygorskite clay, the characteristic diffraction peaks of dolomite and palygorskite completely disappeared in the sample calcinated at 700 °C. Most dolomite disappeared after acid and thermal treatment, and the dehydroxylation of palygorskite at 700 °C resulted in collapse of the pore structure, destruction of the crystal structure and change into the amorphous structure. Because dolomite was not removed completely during acid treatment, reactions between palygorskite and dolomite residue produced new components of MgO and Ca2SiO4 when calcinated at 700 °C. Fig. 5 shows the specific surface areas (SSA) of thermally modified palygorskite clay. The BETSSA of raw palygorskite was 118m2/g. There was no apparent change in SSA when the temperature was increased to 700 °C，because the fiber morphology of paygorskite did not change during the removal of coordinate water and crystal-structure water. When the temperature was increased to 800 °C, the SSA values decreased sharply (from 96 to 33m2/g). The reason was attributed to change of the paygorskite morphology. When calcinations temperature was higher than 800 °C, paygorskite fiber morphology was shrank into sphere，which resulted in tremendous change of the BET-SSA (Chen et al., 2006). Fig. 3 shows that the adsorption capacity of modified palygorskite clay was improved when the modified temperature rose to 700°C and it had a superb adsorption capacity after thermal treatment between 600 °C to 700 °C. However, the BET-SSA decreased with increasing the temperature. This indicates that the enhanced adsorption capacity of modified palygorskite clay was not attributed to the SSA of the particles. The 27Al MAS NMR spectra of raw and modified palygorskite clay are shown in Fig. 6. The spectrum of raw palygorskite (marked sample A in Fig. 6) consists of two resonance peaks. One strong resonance peak at 4.96 ppm indicates the presence of Al in the octahedral coordination, and the other small resonance peak at 56.35 ppm suggests the presence of a little amount of the tetrahedral coordinated Al 1396

(Muller et al., 1981). It shows that Al in palygorskite was mainly in the octahedral coordination, along with a little amount in the tetrahedral coordination (Brindley and Brown, 1980). The spectrum of the modified sample (marked sample B in Fig. 6) shows that the resonance peak at 4.96 ppm disappeared after thermal treatment while the resonance peak at 56 ppm was strengthened. There is a new resonance peak at 18.35 ppm. The 27Al MAS NMR results show that the coordination of Al changed from the octahedral coordination (Al(VI)) to the tetrahedral coordination (Al(IV) during the thermal treatment, the lattice of palygorskite was broken, and a little Al of pentacoordination (Al(V)) appeared in the modified palygorskite clay sample. When the sample was heated above 700 °C and most of the hydroxyls lost, aluminum was activated and exposed at the surface. Activated Al atoms had a high affinity for phosphate anions and provided activity sites for phosphate adsorption. Palygorskite is composed of Si, Mg, Al and a small amount of Fe. The coordination of Al in palygorkite transferred from mainly six-coordinate to mainly four-coordinate, and the structure transferred from the crystal structure to the amorphous structure after calcination at 700 ℃, while the adsorbent had the maximum adsorption capacity. The results above show that the coordination change of Al should be closely related to the increase of phosphate adsorption. Al-OH and Si-OH on the surface of palygorskite were thought to be its active sorption sites (Galan, 1996). Changes of Al coordination may alter the bond length and angle of Al-O and improve the binding activity between AlOH group and phosphate in solutions. Phosphate adsorption activity of palygorskite was increased. 3.3. Column adsorption Fig. 7 shows phosphate adsorption in the column experiment. The accumulative total effluent volume reached 160 L until the effluent P concentration reached 0.025 mg/L (which is GB3838-2002 type II standard for surface water in China). The result shows that the modified palygorskite clay adsorbent had a great capacity for the removal of low concentration phosphate in micropolluted water. The phosphate adsorption capacity of modified palygorskite clay adsorbent was calculated as 5.2 mg/g. Since the adsorbent dosage would be 187.5 g to per ton wastewater treated, the solid liquid ratio for phosphate advanced treatment would be 0.19‰. In addition, palygorskite clay calcinated after phosphate adsorption has a potential application to soil fertilization. Considering the low cost of palygorskite clay in China, this material for advanced phosphorus treatment is economical and effective. Fig. 8 shows the XPS spectrum of phosphorus on modified palygorskite samples.


Fig. 3. Effect of heated temperature for removal efficiency of phosphate (NPAL: not acid treatment; others presented ratio of clay and acid)

Fig. 4. XRD micrographs of modifited and raw palygorskite (A- Raw sample；B- palygorskite clay calcinated at 700)

Fig. 5. Specific surface area change curves of palygorskite calcined at different temperatures

Fig. 6. NMR spectrum of modificated and natural palygorskite (A-Natural palygorskite；B-Modificated palygorskite)

Fig. 7. Phosphate adsorption of column experiment

Fig. 8. XPS spectra of P on modificated palygorskite

There is a 133.6 eV peak of the binding energy of P (2p). Compared with the Handbook of Xray Photoelectron Spectroscopy (Moulder. et. al., 2004), the existence species of phosphorus was H2PO4-. In the meantime, the process of phosphorus

adsorption on modified palygorskite is the coexistence process of physically adsorption and chemistry, which can be certificated by the characterization analyze of the existence of phosphorus.

1397


4. Conclusions Compared with raw palygorskite clay, palygorskite calcinated at 600 – 700 °C exhibited the highest phosphate adsorption capacity. The column experiment revealed that the phosphate adsorption capacity of modified palygorskite clay adsorbent was up to 5.2 mg/g. The highest phosphate adsorption capacity of palygorksite clay after calcined at 600 – 700 °C was attributed to the crystal structure change of palygorskite. After thermal treatment, crystalline palygorskite was distorted due to the dehydroxylation and turned into amorphous. The Al coordination was changed significantly from the octahedral coordination (Al(VI)) to tetrahedral- (Al(IV)) and pentacoordination(Al(V)) during the thermal treatment which enhanced the binding of Al on the surface of palygorskite with phosphate and increased the adsorption capacity of palygorksite. Acknowledgements The authors would like to thank the financial support from the National Natural Science Foundation of China (NSFC: 41102023, 41072036, 41072035, 41210104049). We are grateful to Prof. Wenmin Pang at University Science and Technology China and Dr Yuefei Zhou for the sample test in this study.

References Agyei N.M., Strydom C.A., Potgieter J.H., (2002), The removal of phosphate ions from aqueous solution by fly ash, slag, ordinary Portland cement and related blends, Cement and Concrete Research, 32, 1889– 1897. Akay G., Keskinler. B., Cakici A, Danis U., (1998), Phosphate removal from water by red mud using crossflow microfiltration, Water Research, 32, 717– 726. Al-Futaisi A., Jamrah A., Al-Hanai R., (2007), Aspects of cationic dye molecule adsorption to palygorskite, Desalination, 214, 327–342. Álvarez-Ayuso E., García-Sánchez A., (2007), Removal of cadmium from aqueous solutions by palygorskite, Journal of Hazardous Materials, 147, 594–600. Ayoub G.M., Koopman B., Pandya N., (2001) Iron and aluminum hydroxy (Oxide) coated filter media for low concentration phosphorous removal, Water Environment Research, 73, 478–485, Babatunde A.O., Zhao Y.Q., Zhao X.H., (2010), Alum sludge-based constructed wetland system for enhanced removal of P and OM from wastewater: Concept, design and performance analysis, Bioresource Technology, 101, 6576–6579. Bia G., Pauli C.P., Borgnino L., (2012), The role of Fe(III) modified montmorillonite on fluoride mobility: Adsorption experiments and competition with phosphate, Journal of Environmental Management, 100, 1–9. Blaney L.M., Cinar S., SenGupta A.K., (2007), Hybrid anion exchanger for trace phosphate removal from water and wastewater, Water Research, 41, 1603– 1613.

1398

Bowden L.I., Jarvis A.P., Younger P.L., Johnson K.L., (2009), Phosphorus removal from waste waters using basic oxygen steel slag, Environmental Science and Technology, 43, 2467–2481. Brindley G.W., Brown G., (1980), Crystal Structures of Clay Minerals and Their X-ray Identification, Mineralogical Society, London. Chen T.H., Liu H.B., Li J.H., Chen D., Chang D.Y., Kong D.J., Frost R.L., (2011), Effect of thermal treatment on adsorption–desorption of ammonia and sulfur dioxide on palygorskite: Change of surface acid–alkali properties, Chemical Engineering Journal, 166, 1017– 1021. Chen T.H., Wang J.Y., Shu S.L., Bao T., Gao D.S., Li S.J., (2010), Thermodynamic of adsorbing low concentration phosphate from water on modified palygorskite, Journal of the Chinese Ceramic Society (in Chinese with English abstract), 38, 1816–1819. Chen T.H., Wang J., Qing C.S., Peng S.C., Song Y.X., Guo Y., (2006), Effect of heat treatment on structure, morphology and surface properties of palygorskite, Journal of the Chinese Ceramic Society (in Chinese with English abstract), 34, 1406–1410. Chen T.H., Xu X.C., Yue S.C., (2004) Nanometer Mineralogy and Geochemistry of Palygorskite Clays in the Border of Jiangsu and Anhui Provinces (in Chinese with English abstract), Science Press, Beijing, China. Chen T.H., (2000), A discussion on the adsorption mechanism of palygorskite clay to pollutants in wastewater, Geological Journal of China Universities (in Chinese with English abstract), 6, 265-270. Chen Y.W., Qin B.Q., Teubner, K., Dokulil, M.T., (2003), Long-term dynamics of phytoplankton assemblages: Microcystis-domination in Lake Taihu, a large shallow lake in China, Journal of Plankton Research, 25, 445453. Donnert D., Salecker M., (1999), Elimination of phosphorus from municipal and industrial waste water, Water Science and Technology, 40, 195–202. Fytianos K., Voudriasb E., Raikos N., (1998), Modelling of phosphorus removal from aqueous and wastewater samples using ferric iron, Environmental Pollution, 101, 123–130. Galan E., (1996), Properties and applications of palygorskite-sepiolite clays, Clay Minerals, 31, 443453 Gan F.G., Zhou J.M., Wang H.Y., Du C.W., Chen X.Q., (2009), Removal of phosphate from aqueous solution by thermally treated natural palygorskite, Water Research, 43, 2907–2915. Gao L., Zhou J.M.M., Yang, H., Chen J., (2005), Phosphorus fractions in sediment profiles and their potential contributions to eutrophication in Dianchi Lake, Environmental Geology, 48, 835–844. Genz A., Kornmüller A., Jekel M., (2004), Advanced phosphorus removal from membrane filtrates by adsorption on activated aluminium oxide and granulated ferric hydroxide, Water Research, 38, 3523–3530. Gustafsson J.P., Renman A., Renman G., Poll K., (2008), Phosphate removal by mineral-based sorbents used in filters for small-scale wastewater treatment, Water Research, 42, 189–197. Hano T., Takanashi H., Hirata M., Urano K., Eto S., (1997), Removal of phosphorus from wastewater by activated alumina adsorbent, Water Science and Technology, 35, 39–46.


Hosni K., Moussa S.B., Chachi A., Amor M.B., (2008), The removal of PO43− by calcium hydroxide from synthetic wastewater: optimisation of the operating conditions, Desalination, 223, 337–343. Hrenović, J., Žigovečki G. Ž., Bermanec, V., (2012) Occurrence of sepiolite in Croatia and its application in phosphate removal from wastewater, Applied Clay Science, 1, 64–68. Hsu P.H., (1975), Precipitation of phosphate from solution using aluminum salt, Water Research, 9, 1155–1161. Huang Q.H., Wang Z.J., Wang D.H., Wang C.X., Ma M., Jin X.C., (2005), Origins and mobility of phosphorus forms in the sediments of lakes Taihu and Chaohu, China, Journal of Environmental Science and Health Part A-Toxic/Hazardous Substances & Environmental Engineering, 40, 91–102. Jenkins D., Ferguson J.F., Menar A.B., (1971), Chemical processes for phosphate removal, Water Research, 5, 369–389. Johansson L., Gustafsson J.P., (2000), Phosphate removal using blast furnace slags and opoka-mechanisms, Water Research, 34, 259–265. Jutidamrongphan W., Park K.Y., Dockko S., Choi J.W., Lee S.H., (2012), High removal of phosphate from wastewater using silica sulfate, Environmental Chemistry Letters, 10, 21–28. Karageorgiou K., Paschalis M., Anastassakis G.N., (2007), Removal of phosphate species from solution by adsorption onto calcite used as natural adsorbent, Journal of Hazardous Materials, 139, 447–452. Lee Y., Zimmermann S.G., Kieu A. T., Gunten U., ( 2009), Ferrate (Fe(VI)) Application for municipal wastewater treatment: A novel process for simultaneous micropollutant oxidation and phosphate removal, Environmental Science and Technology, 43, 3831– 3838. Morse G.K., Brett S.W., Guy J.A., Lester J.N., (1998), Review: Phosphorus removal and recovery technologies, Science of Total Environment, 212, 69– 81. Moulder J.F, Stickle W.F, Sobol P.E. (2004), Handbook of X-ray Photoelectron Spectroscopy, Perkin-Elmer Corporation Publisher, Montreal. Muller D., Gessener W., Behrends H.J., Scheler G., (1981), Determination of the aluminum coordination in aluminum-oxygen compounds by solid-state highresolution 27Al NMR, Chemical Physics Letter, 79, 59–62. Murray H.H., (2000), Traditional and new applications for kaolin, smectite, and palygorskite: a general overview, Applied Clay Science, 17, 207–221. Oguz E., (2004), Removal of phosphate from aqueous solution with blast furnace slag, Journal of Hazardous Materials, 114, 131–137. Özacar M., (2003), Adsorption of phosphate from aqueous solution onto alunite, Chemosphere, 51, 321–327. Peleka E.N., Deliyanni E.A., (2009), Adsorptive removal of phosphates from aqueous solutions, Desalination, 245, 357–371. Peng C., Wang S.S., Chen T.H., Jiang S.T., Hang C.H., (2006), Adsorption kinetics of methylene blue from

aqueous solutions onto palygorskite, Acta Geologica Sinica (in Chinese with English abstract), 80, 236– 242. Seida Y., Nakano Y., (2002), Removal of phosphate by layered double hydroxides containing iron, Water Research, 36, 1306–1312. Shin E.W., Han J.S., (2004), Phosphate sdsorption on aluminum-impregnated mesoporous silicates: Surface structure and behavior of sdsorbents, Environmental Science and Technology, 38, 912–917. Tanada S., Kabayama M., Kawasaki N., Sakiyama T., Nakamura T., Araki M., Tamura T., (2003), Removal of phosphate by aluminum oxide hydroxide, Journal of Colloid and Interface Science, 257, 135-140. Wu G.X., Guan Y.T., Zhan X.M., (2011), characteristics of nutrient removal from synthetic wastewater with different organic substrates, Environmental Engineering and Management Journal, 10, 649-654. Xie J.J., Qing C.H., Chen T.H., Guo Y., Pan M., (2007) A comparison of phosphate adsorption capacity between different iron hydroxides/oxides, Acta Petrologica et Mineralogica, 26, 535-538. (in Chinese with English abstract) Xiong J.B., He Z.L., Mahmood Q., Liu D., Yang X., Islam E., (2008), Phosphate removal from solution using steel slag through magnetic separation, Journal of Hazardous Materials, 152, 211–215. Yang Y., Zhao Y.Q., Babatunde A.O., Wang L., Ren X., Han Y., (2006), Characteristics and mechanisms of phosphate adsorption on dewatered alum sludge, Separation and Purification Technology, 51, 193–200. Ye H.P., Chen F.Z., Sheng Y.Q., Sheng G.Y., Fu J.M., (2006), Adsorption of phosphate from aqueous solution onto modified palygorskites, Separation and Purification Technology, 50, 283–290. Zamparas M., Gianni A., Stathi P., Deligiannakis Y., Zacharias I., (2012), Removal of phosphate from natural waters using innovative modified bentonites, Applied Clay Science, 62–63, 101–106. Zelmanov G., Semiat R., (2011), Phosphate removal from water and recovery using iron (Fe+3) oxide/hydroxide nanoparticles-based agglomerates suspension (AggFe) as adsorbent, Environmental Engineering and Management Journal, 10, 1923-1933. Zeng L., Li X.M., Liu J.D., (2004), Adsorptive removal of phosphate from aqueous solutions using iron oxide tailings, Water Research, 38, 1318-1326. Zhang L., Zhou Q., Liu J.Y., Chang N., Wan H., Chen J.H., (2012), Phosphate adsorption on lanthanum hydroxide-doped activated carbon fiber, Chemical Engineering Journal, 185–186, 160–167. Zhao Y.Q., Yue Q.Y., Li Q., Xu X., Yang Z.L., Wang X.J., Gao B.Y., Yu. H., (2012), Characterization of red mud granular adsorbent (RMGA) and its performance on phosphate removal from aqueous solution, Chemical Engineering Journal, 193–194, 161–168. Zhou Y.N., Xing X.H., Liu Z.H., Cui L.W., Yu A.F., Feng Q., Yang H.J., (2008), Enhanced coagulation of ferric chloride aided by tannic acid for phosphorus removal from wastewater, Chemosphere, 72, 290–298.

1399


BIOSACC BIOSORPTION AND BIOACCUMULATION IN THE BIOREMEDIATION OF ENVIRONMENTAL COMPARTMENTS CONTAMINATED WITH PERSISTENT POLLUTANTS Research Grant no. 265/05.10.2011 of IDEI PROGRAMME PN-II-ID-PCE-2011-3-0559 The call emphasizes the need to support and promote fundamental, interdisciplinary and/or exploratory scientific research in Romania so as to assert the international excellence and visibility of scientific research, and to identify and support internationally competitive teams for research and development. In this context, BIOSACC aims to conduct research for the purposes of acquiring advanced knowledge in the field of bioremediation of environmental compartments (air, water, soil) contaminated with persistent pollutants, and offering cost-effective and environmentally friendly technological alternatives to common non-biological approaches. Together with pollution prevention and detection, the elimination of a wide range of pollutants from waters and soils is an absolute requirement for sustainable development. The basis of the proposed research is represented by biosorption and bioaccumulation, which in the present context is the use of non-living and living biomass to reduce pollutant concentration or remove contaminants from environmental compartments, in particular from aqueous solutions (contaminated drinking water and treated wastewater effluents). Project objectives Fundamental objective of the project As an overall objective, BIOSACC will enhance the contribution of environmental biotechnology in the bioremediation of contaminated environmental compartments and improve treatment strategies of water resources and aqueous effluents beyond the state of the art. BIOSACC will exploit the biosorption and/or bioaccumulation potential of non-living and living biomass to remove persistent pollutants from aqueous solutions with various concentrations and thereby to meet the EU quality criteria. The following specific objectives have been established to meet the overall objective: (i) Screening of leading bioremedial processes and adequate regulations and benchmarking, in order to define target values of the intended processes and technologies, target contaminants, biosorbents and bioaccumulators, together with establishing the configuration and scale of the treatment system/bioreactor; (ii) Studies on the biosorption and bioaccumulation of some heavy metals and persistent organic pollutants under various experimental conditions, data evaluation and analysis, modeling the processes, application of new chemical and microbiological tools for exploring the processes; (iii) Enhance the removal efficiency by developing novel adsorbents and bioaccumulators, find the optimal

combination of biosorbent/contaminant/process parameters to form an integrated approach applicable to various water treatment and bioremediation processes; (iv) Provide a rapid cost-effective routine with reliable monitoring opportunities for enhancing the effectiveness and improving the predictability and reliability; (v) Scale-up of laboratory processes at bench scale and simulation of applicability at large scale, provide technical prototypes and know-how for biotreatment; (vi) Integrated assessment and performance validation through cost/benefit analysis, Life Cycle Assessment and risk evaluation, large dissemination of the results. Project description/activities The project is structured in 7 well-integrated work packages (WPs). The work plan consists of two main parts: (i) the scientific work and the technological development will be performed within WP2-WP6. WP3-WP4 aim at exploring the “black box” of biosorption and bioaccumulation, providing in-depth knowledge about mechanisms, kinetic and thermodynamic issues, relations between biomass growing and sorbate binding, various metabolic processes. (ii) This RTD-module will be handled by the project management (WP1) and dissemination (WP7) (Pert Diagram, Fig. 1) WP1 Project management and coordination

WP5

WP3

Models and predictability, application and reliability evaluation and adjustment of BIOSACC systems for large scale applications

Biosorption and bioaccumulation of heavy metals

WP2 Assessment of current treatment approaches. Definition of target compounds, biosorbents and bioaccumulators

WP4

WP6

Biosorption and bioaccumulation of persistent organic pollutants

Integrated assessment and performance validation

WP7 Dissemination activities

Fig. 1. Graphic presentation of the project structure

For more information on the BIOSACC Project please visit: http://biosacc.xhost.ro/. Project Director Professor Maria Gavrilescu, PhD Department of Environmental Engineering and Management “Gheorghe Asachi” Technical University of Iasi, Romania [email protected]


July 2013, Vol.12, No. 7, 1401-1409



TREATMENT OF AN ALKALINE BUTYL RUBBER WASTEWATER BY THE PROCESS OF COAGULATION AND FLOCCULATION HYDROLYSIS ACIDIFICATION - BIOLOGICAL CONTACT OXIDATION - MBR Yan Zhang1, Wei Zheng2, Rui Liu2, Wei Li3, Ying Li2,3, Lujun Chen1,2 1

School of Environment, Tsinghua University, Beijing 100084, P.R. China Zhejiang Provincial Key Laboratory of Water Science and Technology, Department of Environmental Technology and Ecology, Yangtze Delta Region Institute of Tsinghua University, Zhejiang, Jiaxing 314050, P.R. China 3 Zhejiang Envirtecs Water and Wastewater Technology Company, Jiaxing, 314000, P.R. China 2

Abstract Wastewater from butyl rubber production is causing an increasing environmental problem due to its large quantity and complex quality. In this paper, two types of wastewaters generated from rubber synthesis, i.e. alkaline wastewater (AW) and cleaning and flushing wastewater (CFW), were investigated. Coagulation and flocculation were firstly applied to remove suspended solids (SS) from the AW. By single factor experiment, the optimum operational conditions were obtained as pH 8, PAC dosage 40 mg/L, and PAM dosage 8 mg/L, by which the Chemical Oxygen Demand (COD) removal was 31.5±22.4%, and the turbidity of supernatant was 2.7±0.4 NTU. Subsequently, a combined process of hydrolysis acidification (HA) – biological contact oxidation (BCO) – membrane bioreactor (MBR) was designed to treat the mixture of treated AW and raw CFW, according to the biodegradability of the hybrid wastewater. Under the hydraulic retention times (HRTs) of 6h (HA), 4h (BCO), and 10h (MBR), respectively, a COD removal of 88.6± 6.3% was achieved. The organic components of raw water and a series of effluent of each treatment unit were analyzed by GC/MS. Biological acute toxicities of the same samples were measured using luminescent bacterium test. Some transformations of organic components caused by the treatment of HA were revealed. The toxicity of the water was enhanced by HA, and removed by MBR subsequently, which indicated the detoxification of MBR. Key words: biological acute toxicity, biological contact oxidation, butyl rubber wastewater, coagulation, hydrolysis acidification, Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Synthetic rubber (SR) appeared about a century ago as a substitute for the natural rubber which can only be produced in certain regions of the world. New types of rubber are arising rapidly, and the production is increasing significantly owing to expanded demand (Engehausen, 2008). Currently, polybutadiene (BR), styrene butadiene copolymers (SBR), polychloroprene (CR), acrylonitrile-butadiene copolymers (NBR), butyl rubber (IIR), etc. are the 

main types of SR. Generally, the production process of SR comprises polymerization, rubber washing, desiccation, and washing of the reaction tank. Therefore, though the contaminants in wastewaters from diverse rubber production are quite different, the generation and characteristics of the wastewaters are similar. The main wastewater generated from rubber washing and desiccation often is moderately polluted. However, the wastewater generated from reaction tank washing often has extremely high

Author to whom all correspondence should be addressed: E-mail: [email protected], Phone: +861062776686; Fax: +861062797581


concentrations of small rubber particle, salinity, and organic compounds like raw material and polymer. The biodegradability of the wastewater may be moderately fine owing to the monomer residue. But if it was treated insufficiently, organic pollutants of monomer, polymer, initiator, diffusant, emulsifier and solvent may cause certain potential environmental risk. Butyl rubber is a commercial elastomer copolymerized from isobutylene, a small amount of isoprene utilizing methyl chloride (MeCl) as a diluent, and aluminum chloride as an initiator (Gronowski, 2003). This material has been widely used as the chemical raw material of inner tubes, cap sealants, gas pipe coating etc. owing to its desirable physical properties, such as low air permeability, high mechanical strength, excellent heat and steam resistance as well as outstanding chemical resistance (Gronowski, 2003; Pacala et al., 2012; RazzaghiKashani et al., 2007). The global production of butyl rubber is about 1.2 million metric tonnes in 2011. The demand is still growing rapidly. Studies on synthetic rubber wastewater treatment emerged since 1970s. In 1975, Troppe (1975) reported the synthetic rubber wastewater treatment using a complicated process in a series: neutralization, coagulation, flocculation, primary clarification, biological treatment, final clarification, and sludge impoundment. The removals of biochemical oxygen demand (BOD5) and suspended solids (SS) were 84.2% and 85.2%, respectively (Troppe, 1975). In 1983, the total organic carbon (TOC) of a number of wastewaters from synthetic rubber plants was determined by a method of potassium persulfate oxidation (Yakubenok et al., 1983). In 1997, by contacting with ozone-oxygen (2:98, v/v) mixture, the decomposition rate of contaminants in the rubber wastewater was 41-100%, depending on the nature of contaminants (Zimin et al., 1997). Loss et al. (2006) found that the reaction between phosphoric acid (H3PO4) and sodium hydroxide (NaOH) could result in crystallization of large amounts of total dissolved solid (TDS) in the butadiene washing stream from SR industries, and p-tertbutylcatechol (p-TBC) could be recovered from the water. Recently, the application of plasma chemical and plasma catalytic process to degrade pollutants in plastic and rubber wastewater has been studied. The treatment can be associated with photocatalyst which could decrease turbidity and increase biodegradability (Ghezzar et al., 2008). In line with the biochemical treatment processes especially that combined with physicochemical technologies, a styrene-butadiene rubber wastewater was treated by a hydrolysis acidification (HA) - biological contact oxidation (BCO) process in 1997. The system could achieve a COD removal of 87.5%. In 2002, an acrylic rubber wastewater was treated by sedimentation and activated sludge

1402

process, the removals of SS and COD were 77.4%88.7% and 89%-97.8%, respectively (Gao, 2002; Gholikandi et al., 2012). In 2004, the combined process of air floatation - hydrolysis acidification - activity sludge was applied to treat a styrene-butadiene rubber wastewater with COD and BOD5 removal of 83.5%87.9% and 93.6%-95.2%, respectively (Yang et al., 2004). In 2010, a certain synthetic rubber wastewater has been treated by the combined process of hydraulic circulating clarifier and membrane bioreactor (MBR) (Qiu et al., 2010). Though the production process of IIR is similar to that of other SR, the sources of IIR wastewater are not identified, and thereby its quality remains unclear. This forms the basis of the current study. According to our preliminary study, alkaline wastewater (AW) from alkaline washing of IIR productive reactor, which is the most serious polluted wastewater, contained high concentrations of organic pollutants, SS, and TDS. These should be treated using an integrated process. After comparison of different available technologies, coagulation and flocculation, as a pretreatment of biological process, was adopted to remove solid particles. It was followed a biochemical process combined of HA - BCO - MBR to remove the organic pollution. The optimum operational parameters of coagulation & flocculation, performance of biochemical treatment, transformation of organic componentsand biological acute toxicity were investigated. The purpose of this study is to explore an effective and biosafety way to treat this kind of industrial wastewater. 2. Materials and methods 2.1. Materials 2.1.1. Wastewaters Two types of wastewater, AW and a cleaning and flushing wastewater (CFW) generated from devices and floor washing, from rubber synthesis in a butyl rubber factory in Zhejiang Province, China were used in this study. The characteristics of the two types of wastewater are shown in Table 1. 2.1.2. Reactors The treatment system of HA-BCO-MBR was made of plexiglass and consisted of three bench-scale reactors. The sizes (diameter×height) of HA, BCO, and MBR were Φ0.15×1.0 m, Φ0.10×1.2 m, and Φ0.25×0.6 m, respectively, with the working volumes of 8.2 L, 5.5 L, and 14.0 L, respectively. Both HA and BCO were filled with 8 pieces of semisoft fibrous-plastic media. The area of membrane fixed in the MBR was 0.2 m2 while the flux was 12.5 l/(m2·h). The schematic diagram of the system is shown in Fig. 1.

Treatment of an alkaline butyl rubber wastewater

Table 1. Wastewater characteristics Wastewater

pHb

AW CFW

>14.0 8.2±1.2

CODb （mg/L） 8980.6±7210.2 133.1±142.7

BOD5c （mg/L） 3569.2±2790.0 86.1±14.2

SSc (mg/L) 523.2±444.2 -a

TDSc （g/L） 28.73±23.09 0.73±0.05

Turbidityc （NTU） 537.0±65.0 20.2±29.8

Colorityc (Degree) 749±58 -a

a. not measured; b. pH and COD were measured for more than 50 times; c. BOD5, SS, TDS, turbidity, and colority were measured for more than 10 times.

Fig. 1. Schematic diagram of the HA-BCO-MBR system 1. storage tank; 2. peristaltic pump; 3. HA; 4. 5. semi-soft fibrous-plastic media; 6. BCO; 7. air compressor 8. MBR; 9. membrane module; 10. peristaltic pump

2.2. Methods 2.2.1. Biodegradability test Biodegradability of the wastewater was determined by comparing the oxygen consumption to endogenous respiration of the cultivated sludge. The activity sludge was aerated for 24 h, and then washed with normal saline for three times before use. The two types of wastewaters were initially neutralized to pH 7, then inoculated with the pretreated activity sludgeand cultivated in a Warburg respirometer (KEDA SKW-3, China) to test the biodegradability. Deionized water substituting the wastewater was adopted to determine oxygen consumption of endogenous respiration as the blank control. 2.2.2. Coagulation & Flocculation Polyaluminium chloride (PAC) and polyacrylamid (PAM) were used as coagulant and flocculant. The experiment was conducted in beakers using magnetic stirrers (Honghua HJ-6, China). The factors of pH and dosages of PAC and PAM, were individually optimized by single factor experiments. pH was optimized under the condition of PAC dosage, 50 mg/L and PAM dosage, 10 mg/L. Then, dosage of PAC was optimized with the optimum pH and PAM dosage, 10 mg/L. Finally, dosage of PAM was optimized under the optimum pH and PAC dosage. The rapid mixing time after dosing with PAC was 30 s while the slow stirring lasted 10 min after dosing with PAM.

2.2.3. Operation parameters of the process The HRTs of HA, BCO, MBR were 6 h, 4 h, and 10 h, respectively. The dissolved oxygen (DO) concentrations of BCO and MBR were around 3 mg/L and 4 mg/L, respectively. The influent of the system was prepared by mixing the pretreated AW with the CFW for the purpose of eliminating the inhibition of high TDS in the AW. At the beginning, the proportion of AW was controlled at a low level of 20%, for weakening the inhibition. After 15 days operation, the ratio of AW and CFW was adjusted to 1.5:4, and then on the 40th day, it was further adjusted to a relatively high level of 1:2 to enhance the treatment velocity of AW. The concentrations of ammonia and phosphate of the mixed wastewater were only 0-2 mg/L, which could not meet the requirements of the biochemical treatment. As such, extra amounts of urea and KH2PO4 were added to the influent to ensure the ratio of BOD: N: P to be around 100: 5: 1. 2.2.4. Wastewater quality and organic components analysis COD, BOD5, total nitrogen (TN), total phosphorus (TP), pH, SS, TDS, and turbidity were analyzed using standard methods (The former State Environmental Protection Administration, 2002). Colority was determined by a colorimeter (SD-9011, Shanghai Xinrui, China) using platinum-cobalt colorimetry. Considering low boiling point and volatility of the materials involved in butyl rubber manufacturing process, gas chromatography - mass 1403


3. Results and discussion

condition. It was decreased with the increase of pH under alkaline conditions of pH > 8. The colority of supernatant also achieved the minimum value when pH was 8. It can be concluded that the optimum pH is 8. This is consistent with the requirement of neutralization before biochemical treatment. Under the conditions at pH 8 and PAM dosage of 10 mg/L, the dosage of PAC (10-80 mg/L) affected the treatment results greatly as shown in Fig. 3. The COD removal achieved the maximum, and the colority of supernatant reaches minimum, simultaneously, under the optimum dosage of PAC, 40 mg/L. Under the conditions at pH 8 mg/l and PAC dosage of 40 mg/l, the dosage of PAM (2-10 mg/l) influenced the treatment results markedly as Fig. 4 shown. The colority of supernatant decreased rapidly with the increase of PAM dosage. The colority achieved the minimum value of 31 at the dosage of 8 mg/L. Then it was rebounded to a higher value of 55 at the dosage of 10 mg/L. The COD removal reached the highest rate at the dosage of 8 mg/L. Therefore, the optimum dosage of PAM was determined as 8 mg/L. Operated in a sequencing batch mode under the above optimum conditions, the coagulation & flocculation played a important role of pretreatment of the HA-CO-MBR process. The average COD removal was 31.5±22.4%. The turbidity of supernatant was 2.7±0.4 NTU. Actually, PAC and PAM have been extensively adopted in coagulation and flocculation (Chaudhari et al., 2010; Kushwaha et al., 2010; Yang et al., 2010). There are different forms of aluminum presented in PAC including Al3+, Al(OH)2+, Al2(OH)42+, and Al(13) polymer or higher polymers. During the treatment, pH can affect the proportions of these aluminum ions and polymers which can influence the treatment performance. In a slightly alkaline environment, polymeric species, as the main form, can be adsorbed onto the surface of colloidal particles, resulting in the promotion of colloid aggregation and sedimentation. However, Al(OH)3, which is negative for coagulation efficiency, can also increase sharply in the alkaline solution, and finally become the main form when pH achieves a certain value. In contrast, in an acid environment, Al3+, existing as the main form, can reduce the coagulation efficiency owing to its disadvantage for adsorption adhesion, bridges, cross-linking etc. (Zheng et al., 2011). The optimum pH of 8 in this study is a reflection of the principle mentioned above. The differences of COD removals among different batches were relatively distinct, that is because of deferent concentration of influent SS. The satisfactory treatment performance indicated that coagulation & flocculation was an appropriate pretreatment for AW.

3.1. The performance of coagulation & flocculation

3.2. The biodegradability of wastewater

The effects of pH on COD removal and colority of supernatant were investigated and the results are shown in Fig. 2. The removal efficiency of COD increased with the increase of pH under acidic

Warburg respirometer was employed in this study to explore the wastewater biodegradability. This instrument can precisely measure the sludge oxygen uptake (SOU) of the activated sludge

spectrometry (GC/MS) using manual headspace injection was adopted to determine the organic components in the wastewater. The GC-MS is consisted of GC 6890 and MS 5973 manufactured by Agilent Inc., USA. Headspace vial filled with 50 ml wastewater was heated in water bath at 60oC for 30 min, then the gas in the vial was extracted to inject immediately. The split vent was closed during the injection. A capillary column HP-INNOWAX, 30 m×0.25 mm ID with 0.25 um film thickness, was used for separation using helium carrier gas programmed at a constant flow of 1 ml/min. The column oven was programmed to hold at 35 oC for 3 min, increase to 100 oC at 6 oC /min and hold for 2 min, then increase to 155 oC at 4 oC/min and hold for a further 10 min. The transfer line to the mass spectrometer was heated to 160 oC and the ion source was operated at 230 oC. In MS mode, the scan range was 30-350 atomic mass unit (amu). The compounds were identified by comparison of the recorded spectral data with the corresponding spectra in the NIST 2008 library (Gaithersburg, MD, USA) with all computer spectral matches checked manually. 2.2.5. Biological acute toxicity The luminescent bacterium test is a rapid and effective method to evaluate the acute toxicity of wastewater by determining the inhibition of light output of luminescent bacterium (Fernandez-Alba et al., 2001; Ye et al., 2011). The acute toxicity tests of raw wastewater and a series of effluent of each treatment unit were conducted according to GR/T 15441-1995 using a BHP9511 water toxicity analyzer (Beijing HAMAMATSU, China). A luminescent bacterium, Photobacterium phosphorem T3, was used in the test. Wastewater was diluted to a series of test samples with concentration gradient. 2 mL of a sample was mixed with 1 mL of the bacterial suspension in a test tube. After 15 min reaction, the relative light intensity (LI) was measured by the instrument. The luminescence inhibition ratio (LIR) was calculated according to the Eq. (1): LIR (%) 

LI ref  LI s LI ref

 100

(1)

where: LI is the light intensity of the luminescence emitted from the luminescent bacteria. The subscripts ref and s represent the reference and sample, respectively.

1404


inoculated in wastewater and has been utilized in wastewater biodegradability analysis by many researchers (Campin, 1980; Olanrewaju, 1988; Yang et al., 2001). The results of this study are shown in Fig. 5. The oxygen uptake of the sludge inoculated in AW was much higher than that of blank control, which means that the AW contained biodegradable fraction. It did not exhibit biological inhibitory. At least, inhibitory did not significantly affect biodegradation in AW. Therefore, biological treatment could be applied as the main process. However, the OUR of CFW was almost coincident with that of blank control, indicating that the CFW was not biodegradable, but without inhibition on the microorganisms in the sludge. 3.3. The performance of HA It is believed that hydrolysis acidification can decompose particulates or refractory organic matters to soluble and biodegradable substances and change the concentration of SCOD, resulting in the improvement of wastewater biodegradability (Chen et al., 2007). The biodegradability of the mixed wastewater would be influence by the CFW, though the biodegradability of AW was good. Therefore, HA was designed as the first step in the process to improve the overall biodegradability.

The BOD5/COD (B/C) values of the HA reactor’s influent and effluent were measured for indicating the biodegradability changes. The results are shown in Fig. 6. As the proportion of AW in the mixed wastewater was improved, the organic loading of influent increased, and the values of B/C rose from 0.26 to 0.47. However, the B/C values of effluent remained more stable than that of influent, and the values were 0.440.56. Under the low organic loading, the B/C was improved from 0.26 to 0.56 on 10th day. However, it was improved from 0.47 to 0.48 under a higher organic loading on 49th day. It indicates that the HA presents an effective function on biodegradability improvement of wastewater under the low organic loading, especially when the ratio of AW and CFW is 1:4. The reasons for this phenomenon may involve two aspects. On one hand, improvement of the proportion of AW increases the proportion of the organic loading which is biodegradability. On the other hand, the loading of toxic or harmful substance may be increased too. 3.4. COD removal of the biological treatment system The results shown in Fig. 7 illustrate the COD removal during the operation of HA-CO-MBR system.

Fig. 2. Effect of pH on the performance of coagulation and flocculation

Fig. 3. Effect of PAC on the performance of coagulation and flocculation

Fig. 4. Effect of PAM on the performance of coagulation and flocculation

Fig. 5. The biodegradability test

1405


Fig. 6. The biodegradability improvement by HA

The COD of the influent was 438-1756 mg/L with the mixed ratio of AW/CFW at 1:4-1:2. Although HA functioned mainly in the biodegradability improvement of the mixed wastewater, the average COD removal by HA was 10.8±12.9%, which was as expected and similar to other studies (Zhang et al., 2010; Lin et al., 2011). The COD removals by the BCO and MBR were 59.3±21.3% and 55.0±18.1%, respectively. When the organic loading increased, the COD removal of the whole process was decreased to 50% in the loading adjustment phase. But the system efficiency recovered rapidly as it adjusted to the new loading. Excluding the starting up period and shock loading adjustment, the COD removal by the HA-BCO-MBR system was 88.6±6.3%. BCO is an efficient process which is widely applied in industrial wastewater treatment. Though the HRT of BCO (4 h) in this study was much shorter than others (Cao et al., 2011; Lu et al., 2011), COD removal of 59.3% was only slightly less. The volumetric loading of BCO was 5.0±1.5 kgCOD/(m3·d). Owing to the higher efficiency of BCO, the influent COD of MBR was controlled at lower level (335.8±192.4 mg/L), and the volumetric loading was 0.8±0.5 kgCOD/(m3·d). MBR presented a stable performance as the COD of effluent was 105.0±53.3 mg/L, which met the discharge requirement of grade I (120 mg/L) for petrochemical industry in China Integrated Wastewater Discharge Standard (GB 8978-1996). 3.5. Transformation of organic components The organic components of influent and effluent of each reactor were measured by GC/MS to analyze the biotransformation as the wastewater was treated progressively by the system. The results are shown in Fig. 8. Four main organic compounds were detected in the influent, including 1aziridineethanamine, methyl tert-butyl ether

1406

Fig. 7. COD change in the treatment system and the total removal rate “■”: influent, “○”: effluent of HA, “▲”: effluent of BCO, “☆”: effluent of MBR, “◆”: COD removal of the system

(MTBE), pentamethylbenzene, and hexamethylbenzene. Part of MTBE was hydrolyzed to metheanol and isobutanol by the HA. However, there was no volatile fatty acid (VFA) determined, inferring that acidification was not significant. After treatment in the BCO, methanol generated in the HA reactor was completely removed. In the effluent of MBR, 1-aziridineethanamine was completely removed by MBR. However, a new compound, chloromethane, was found in the MBR effluent. Its peak occurred at 0.992 min, which was very close to that of 1-aziridineethanamine at 0.981 min. When 1aziridineethanamine existed, the peak of chloromethane might hide behind. Chloromethane could not be determined until 1-aziridineethanamine was removed by MBR. Pentamethylbenzene and hexamethylbenzene didn’t transform in the system, except reducing in quantity. Though the measurement of organic components was not quantitative determined, the linear relationship between the total-peak area of all organic components and the COD concentrations of the corresponding wastewater can be clearly identified in Fig. 9, where the r2 of the fitting equation is 0.9857, indicating that the reduction trend of main organic compounds is consistent with COD. 3.6. Biological acute toxicity of wastewater Significant changes of biological toxicity of the wastewater were found when it was treated in the sequence of HA, BCO, and MBR. The LIR of influent, effluent of HA, BCO, and MBR were 2.60%, 98.92%, 98.59%, and 0.00%, respectively. The LIR of influent (2.60%) indicated that there was nearly no acute toxicity in the wastewater. However, after it was treated by HA, the LIR was increased to 98.92%. This phenomenon was different from other studies, which demonstrated that biological treatment was the main process for toxicity removal (Li et al., 2009).


The possible reason might be that certain substances were generated in HA reactor which restrained the illumination of the sensitive luminescent bacterium. However, the biofilm in HA and BCO reactor can immune the toxicity. Unfortunately, it was hard to approach the precise substance in the current results of organic components analysis. BCO seems to have no significant influence on the wastewater toxicity. The LIR of its effluent was 98.59%. The dose-response curves of effluent of HA and BCO were detected and the results are shown in Fig. 10. The EC50 were calculated by statistical software SPSS. The values were 21.39% and 19.61%, meeting the standard of Grade I (1.0W/mL ultrasound with a maximum voltage of 491 mV, power density of 10.19 W/m3 and 7-d TCOD removal rate of 62.5%. For stable power generation and effective removal of the waste in the MFC system, the raw waste required a certain dilution and the anode solution should be adjusted to improve system buffering capacity. Acknowledgements This research was supported by “National Natural Science Foundation of China” (No. 21077012) and “IndustryAcademia-Research program of Guangdong Province” (No.20090912).

References APHA, (1998), Standards Methods for the Examination of Water and Wastewater, 20th ed. American Public Health Association, Washington, DC, USA. Behera M., Jana P.S., More T.T., Ghangrekar M.M., (2010), Rice mill wastewater treatment in microbial fuel cells fabricated using proton exchange membrane and earthen pot at different pH, Bioelectrochemistry, 79, 228-233. Ciubota-Rosie C., Gavrilescu M., Macoveanu M., (2008), biomass - an important renewable source of energy in

Composite vegetable degradation and electricity generation in microbial fuel cell with ultrasonic pretreatment

Romania, Environmental Engineering and Management Journal, 7, 559-568. Duteanu N.M., Ghangrekar M.M., Erable B., Scott K., (2010), microbial fuel cells - an option for wastewater treatment, Environmental Engineering and Management Journal, 9, 1069-1087. Feng Y.J., Wang X., Logan B.E., Lee H., (2008), Brewery wastewater treatment using air-cathode microbial fuel cells, Applied Microbiology and Biotechnology, 78, 873-880. Frolund B., Palmgren R., Keiding K., Nielsen, P.H., (1996), Extraction of extracellular polymers from activated sludge using a cation exchange resin, Water Research, 30, 1749-1758. Fortună M.E., Simion I.M., Gavrilescu M., (2012), Assessment of sustainability based on LCA – case of woody biomass, Cellulose Chemistry and Technology, 46, 493-510. Greenman J., Galvez A., Giusti L., Ieropoulos I., (2009), Electricity from landfill leachate using microbial fuel cells: comparison with a biological aerated filter, Enzyme Microbiology and Technology, 44, 112-119. He Z., Shao H.B., Angenent L.T., (2007), Increased power production from a sediment microbial fuel cell with a rotating cathode, Biosensors and Bioelectronics, 22, 3252-3255. Huang L., Zeng R.J., Angelidaki I., (2008), Electricity production from xylose using a mediator-less microbial fuel cell, Bioresource Technology, 99, 41784184. Jiang J., Zhao Q., Wei L., Wang K., Lee D-J., (2011), Degradation and characteristic changes of organic matter in sewage sludge using microbial fuel cell with ultrasound pretreatment, Bioresource Technology, 102, 272-277. Khanal S.K., Grewell D., Sung S., van Leeuwen J., (2007), Ultrasound applications in high solids pretreatment: a review, Critical Reviews in Environmental Science and Technology, 37, 1-37. Li Z., Yao L., Kong L., Liu H., (2008), Electricity generation using a baffled microbial fuel cell convenient for stacking, Bioresource Technology, 99, 1650-1655. Liu H., Cheng S., Logan B.E., (2005), Production of electricity from acetate or butyrate in a single chamber microbial fuel cell. Environmental Science and Technology, 39, 658-662. Liu H., Ramnarayanan R., Logan B.E., (2004), Production of electricity during wastewater treatment using a single chamber microbial fuel cell, Environmental Science and Technology, 38, 2281-2285. Logan B.E., Hamelers B., Rozendal R., Schroder U., Keller J., Freguia S., Aelterman P., Verstraete W., Rabaey K., (2006), Microbial fuel cells: methodology and technology, Environmental Science and Technology, 40 , 5181-5192.

Lovley D.R., (2006), Bug juice: harvesting electricity with microorganisms, Nature Reviews Microbiology, 4, 497-508. Min, B., Kim, J.R., Oh, S.E., Regan, M.J., Logan, B.E., (2005), Electricity generation from swine wastewater using microbial fuel cells, Water Research, 39 , 49614968. Oh S.E., Logan B.E., (2005), Hydrogen and electricity production from a food processing wastewater using microbial fuel cell technologies, Water Research, 39, 4673-4682. Rabaey K., Verstraete W., (2005). Microbial fuel cells: novel biotechnology for energy generation, Trends in Biotechnology, 23, 291-298. Raghavulu, S.V., Mohan S.V., Goud R.K., Sarma P.N., (2009), Effect of anodic pH microenvironment on microbial fuel cell (MFC) performance in concurrence with aerated and ferricyanide catholytes, Electrochemistry Communications, 11, 371-375. Scott K., Murano C., (2007), A study of a microbial fuel cell battery using manure sludge waste, Journal of Chemical Technology and Biotechnology, 82, 809817. Shimoyama T., Komukai S., Yamazawa A., Ueno Y., Logan B.E., Watanabe K., (2008), Electricity generation from model organic wastewater in a cassette-electrode microbial fuel cell, Applied Microbiology and Biotechnology, 80, 325-330. Tiehm A., Nickel K., Nies U., (1997), The use of ultrasound to accelerate the anaerobic digestion of sewage sludge, Water Science and Technology, 36, 121-128. Venkata Mohan S., Chandrasekhar K., (2011), Solid phase microbial fuel cell (SMFC) for harnessing bioelectricity from composite food waste fermentation: Influence of electrode assembly and buffering capacity, Bioresource Technology, 102,7077-7085. Venkata Mohan S., Mohanakrishna G., Sarma P.N., (2010), Composite vegetable waste as renewable resource for bioelectricity generation through non-catalyzed openair cathode microbial fuel cell, Bioresource Technology,101, 970-976. Venkata Mohan S., Mohanakrishna G., Srikanth S., Sarma P.N., (2008), Harnessing of bioelectricity in microbial fuel cell (MFC) employing aerated cathode through anaerobic treatment of chemical wastewater using selectively enriched hydrogen producing mixed consortia, Fuel, 87, 2667-2677. Yu G.H., He P.J., Shao L.M., Zhu Y.S., (2008), Extracellular proteins, polysaccharides and enzymes impact on sludge aerobic digestion after ultrasonic pretreatment, Water Research, 42, 1925-1934.

1427


INTERDISCIPLINARY TRAINING AND RESEARCH PLATFORM HIGH PERFORMANCE MULTIFUNCTIONAL POLYMERIC MATERIALS FOR MEDICINE, PHARMACY, MICROELECTRONICS, ENERGY/INFORMATION STORAGE, ENVIRONMENTAL PROTECTION The Platform aims to develop training and interdisciplinary research in high-performance multifunctional polymeric materials. The nucleus of the Platform is based on the Center of Excellence POLYMERS, officially accredited by CNCSIS (7.06.2003), center acting within the “Gheorghe Asachi” Technical University of Iasi. The Platform will be integrated in national/European networks and will ensure the training and improvement of human resources through high education and research, will enhance the research performance and the visibility of Romania, will contribute to Romanian high education and research integration in European Education Area and European Research Area, to the development of the knowledge-based society and will increase the socio-economic impact of research. To ensure the success of the Project, a set of specific objectives has been defined:  New educational programmes, oriented towards European priorities, able to ensure highly qualified human resources and to integrate them into the knowledge-based modern society  New contents, forms and methods of training, specific for the development of education and research in multifunctional polymeric materials and in agreement with Lisbon Agenda and with Bologna Process, as well as with Romanian priorities  Elaboration and implementation of interdisciplinary programmes of training (master, doctoral, post-doc)  Consolidation of excellence in research in the field of high performance multifunctional materials by promoting interdisciplinary programmes and by attracting the most talented graduates – from Romania and abroad – for PhD and post-doc studies  Extension and consolidation of the research infrastructure (hard equipment) of the Platform, to improve the training and research process, in order to increase Platform competitiveness in accessing national (CNCSIS, CEEX, PNCDI 2) and international (FP7, NATO, NSF etc.) programmes and the efficiency in answering the requirements of the regional, national and European economic areas  Strengthening the scientific cooperation with academic and economic partners at national and European level  Promoting the exchange of information and communication between the academic and socioeconomic environments, to consolidate the knowledge-based society and to accelerate the integration of Romania into the European Union. The Project will develop (i) education activities through (i-a) master studies (two directions are proposed – Biomaterials – addressed to graduates of chemistry, chemical engineering, medical bioengineering, biology, medicine, pharmacy – and Multifunctional Materials for Advanced Technologies, addressed to graduates of

chemistry, chemical engineering, medical bioengineering, physics, electronics and electrical engineering, civil engineering, environment protection; both master programmes will be in Romanian and/or English), (i-b) doctoral studies with a pronounced interdisciplinary character and implemented within the “co-tutelle” system, (i-c) post-doc studies (financed from other programmes), and (ii) research activities developed within five programmes, i.e., (ii-a) Biomaterials. Polymer-drug Systems with Controlled and Targeted Release (polymerdrug conjugates, diffusional systems, drug inclusion in polymeric micro- or nanoparticles), (ii-b) Smart Multifunctional Polymeric Materials (molecular imprinting, diagnostics and bioseparation, nanocapsules and nanostructured membranes via core-shell particles, smart hydrogels and nanostructured gels, biomimetic polymeric networks, nanofabrication), (ii-c) Motile Molecular Systems (hybrid and organic polymers for biology, microelectronics, nanorobotics and energy/information storage), (ii-d) Liquid Crystal Heteroorganic and Organic Compounds (liquid crystals for displays, opto-electronic devices, ferro-electric liquid crystals), (ii-e) Molecular Modeling and Artificial Intelligence (conformational analysis and simulation of properties, neuronal networks, fuzzy systems). All planned activities and actions are based on a deep analysis of the tendencies in the interdisciplinary education and research, on the requirements of the national and European market. Most of Platform budget is dedicated to the serious improving of the research infrastructure (hard equipments). Additional funding and expertise will be obtained through the facilities offered by the “Gh. Asachi” Technical University of Iasi, the infrastructure and human resources of the POLYMER Centre of Excellence, through the facilities offered by the traditional national and European partners of the Platform. Platform sustainability will be ensured by different funding attracting activities – training of specialists from SMSs, consulting activities, national and international grants, the RENAR accredited laboratories, the Technology Transfer Center and the Innovation Relay Centre established within the Platform, the specific activities to be performed within the Science and Technology Park in Iasi. The benefits of the Platform will cover the whole high education and research environment in Iasi and in the North-Eastern Region of Romania and all Platform partners – both academic and economic.

Constanţa Ibănescu Department of Natural and Synthetic Polymers Faculty of Chemical Engineering and Environmental Protection, “Gheorghe Asachi” Technical University of Iasi, Romania


July 2013, Vol.12, No. 7, 1429-1436



EFFECTS OF COD:SULFATE RATIO ON SULFATE REMOVAL FROM OIL SHALE RETORT WATER USING MICROBIAL FUEL CELLS Ho Il Park, Lian-Shin Lin West Virginia University, Department of Civil and Environmental Engineering, Morgantown, WV 26506, USA

Abstract This paper reports the effects of COD:sulfate ratio on sulfate removal and electricity generation from oil shale retort water treatment using microbial fuel cells (MFCs). Field-collected retort water was augmented with organics to obtain a range of initial COD:SO42- ratios (0.5:1 to 2:1), and treated in two different MFC designs (tubular and two-chambered). The two-chambered MFCs exhibited COD and sulfate removal 1 to 2 orders of magnitude higher than those of the tubular MFCs. The tubular MFCs did not exhibit a significant dependence of sulfate removal on the COD:SO42- ratio while the two-chambered MFCs showed a positive trend. The tubular MFCs generated a maximum power density of 19 mW/m2 (COD:SO42- = 1.5:1), and the twochambered MFCs produced 120 mW/m2 (2:1). The results suggest that organic carbon loading to the MFCs should be determined based on the sulfate concentration and reactor design to achieve optimal sulfate removal and electric power output. Key words: COD:sulfate ratio, microbial fuel cells, oil shale retort water, sulfate removal Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Projected long-term decline in the supply of fossil fuels such as coal and petroleum and their associated environmental pollution have led to the exploration of alternative energy sources (Jeffrey et al., 2003). Oil shale resources worldwide of about 2.6 trillion barrels are located within the United States (Speight, 2008). This represents a long-term energy source that can help meet future energy needs (Dyni, 2006). Extraction and processing of oil shale involves significant amounts of water and its availability continues to be viewed as a major constraint on large-scale oil shale development (Bartis et al., 2005). There are a number of waterconsuming operations at oil shale plants, such as spent shale moistening, revegetation, dust control, cooling water make-up, and boiler feed water (Petersen, 1981). To minimize water use, reuse and recycling of the generated wastewater on site is a 

potentially effective water management strategy for oil shale extraction and plant operation (Talve and Riipulk, 2001). In particular, water produced from the retorting process of oil shale extraction (retort water) requires treatment to prevent environmental pollution. Retort water typically has a pH ranging from 8 to 9 and contains a wide range of inorganic and organic constituents (Fox, 1981). Major constituents include metals, ammonia/ammonium, carbonate, sulfate, chloride, and organic carbons (Ellis et al., 1995). In particular, sulfate is consistently reported to be present at high levels in retort waters, and its removal from retort water is the focus of this study. Methods proposed for treating retort water include physical methods such as air stripping, flocculation/coagulation, adsorption, reverse osmosis wet air oxidation (Zhang et al., 2006), chemical methods such as chemical oxidation (Sebastian et al., 1996), and biodegradation (Clarke et al., 2005).


Park and Lin/Environmental Engineering and Management Journal 12 (2013), 7, 1429-1436

These methods have been used to remove targeted constituents in retort water. Sulfate removal using biological processes can be achieved by converting sulfate to biogenic sulfide (S2-) using sulfate-reducing bacteria (SRB) (Battaglia et al., 2000; Costa et al., 2008). Several research groups have recently explored sulfate and sulfide removal using microbial fuel cells for electricity generation (Rabaey et al., 2005; Zhao et al., 2008). The chemical fate of sulfur and the associated electrochemical and microbial processes in MFCs can be quite complicated. For example, sulfate was reported to be reduced to sulfide by SRB and reoxidized to sulfate on anode electrodes in an MFC setting (Habermann and Pommer, 1991). Rabaey et al. (2006) provided evidence for sulfide reoxidization to elemental sulfur and other crystallized deposits on granular graphite anode surfaces, following bacterial sulfate reduction. The authors also suggested that hydrogen sulfide (H2S) loss to the atmosphere via volatilization, and the formation of polysulfides, were other possible fates of sulfur. Zhao et al. (2008) reported on sulfate reduction to sulfide in an MFC using Desulfovibrio desulfuricans and the formation of elemental sulfur on anode surfaces. Dutta et al. (2009) found that aqueous sulfide and electrodeposited sulfur could form a mediating system for acetate oxidation in the anode chamber of a bioelectrochemical cell, where biogenic sulfide produced from the elemental sulfur could be reoxidized on anode surface when a favorable anode potential was maintained. At present, the effects of retort water organic carbon levels on sulfate reduction in MFC settings are poorly understood. This study examined the effectiveness of two different configurations of MFCs used for sulfate removal from field-collected retort water under a range of carbon:sulfate ratio. The retort water was augmented with organic carbon (using glucose as a model) to investigate the effects of COD:SO42- ratio on sulfate removal. Complimentary analyses and Coulombic efficiency calculations were performed to help determine the possible fate of sulfur in the reactors. Electrical output of the MFCs was measured and factors that affected electricity generation from the retort water treatment were evaluated. 2. Material and methods 2.1. Retort water A retort water sample (20 L, pH 8.6) was collected at an in-situ oil shale retorting site in the Green River Basin near Rock Springs, Wyoming. The retort water was a mixture of water from 5 wells that had undergone air sparging treatment to remove volatile organic compounds and, as a result, contained a relatively low initial concentration of organics (COD = 220 mg/L). A high concentration of sulfate (2,907 SO42mg/L) was present in the retort water. The water was

1430

sterilized in an autoclave at 121C for 15 min before it was used in the experiments. In the laboratory, the retort water was supplemented with glucose (Thermo Fisher Scientific, Inc.) as a model compound of carbon source to obtain a range of COD:SO42- ratio. 2.2. Microbial fuel cells Two different designs of MFCs were tested for their ability to remove sulfate from retort water. The first design was a tubular reactor (0.6 L) containing three vertically aligned anode electrodes (surface area = 6.0 × 10 cm for each) and two cathode electrodes (diameter 6.0 cm) placed around the top of the aqueous solution. A perforated silicone tube was wrapped between the two cathodes to provide oxygen from air bubbles using an air-pump (Aqua Culture, Eiko Electric Products Corp., USA). In the second design, the two-chambered MFC reactor consisted of an anode chamber (0.02 L) and a cathode chamber (0.02 L) separated by a cation exchange membrane (Nafion-112, Dupont, USA). Both chambers were sealed with a gasket and each contained an electrode (surface area = 2 × 10 cm). A peristaltic pump was used to recirculate aerated phosphate buffer solution (34.84 g/L of K2HPO4, and 27.22 g/L of KH2PO4, pH 7.0) through the cathode chamber to provide oxygen to the cathode. A graphite felt material (Electrosynthesis Co., Lancaster, NY) was used for all of the electrodes. The anode(s) and cathode(s) in each MFC were connected by an external wire loaded with a tunable resistor. Electrons harvested by the anode electrode(s) were routed via the wire to the cathode(s), where oxygen was reduced to water. Two reactor units of each design were set up for retort water treatment: the tubular MFCs are referred to as MFC-1 and MFC-2, and the two-chambered are referred to as MFC-3 and MFC-4. These designs are schematically illustrated in Fig. 1. 2.3. Microbial inoculation and enrichment A sewage sludge sample was collected from an anaerobic digester at a local wastewater treatment plant in Star City, West Virginia. A mixture of the anaerobic sludge (0.1 L), retort water (1.5 L), glucose (1 g), and a phosphate buffer solution (50 mL, 1M, pH 7.0) was used to inoculate the anodes of the MFCs. The inoculation was followed by a threeweek enrichment of the microbial population, during which nitrogen gas was bubbled into the anode chambers to maintain anaerobic conditions. During the enrichment period, glucose was added to the anode chambers to enhance the microbial populations and the cathode electrodes were provided with oxygen by air bubbles or aerated water as described previously. After the microbial enrichment period, the aqueous solutions were drained and the reactors were replenished with fresh retort water augmented with glucose for sulfate removal experiments.

Effects of COD: sulfate ratio on sulfate removal from oilshale retort water using microbial fuel cell

Fig. 1. Schematic illustrations of the tubular MFC and two-chambered MFC

2.4. Sulfate and COD removal experiments Retort water treatment was operated in a fedbatch mode. To investigate the effects of organic carbon on sulfate removal, different amounts of glucose were added to the retort water to obtain a range of initial COD:SO42- mass ratios between 0.5:1 and 2:1. During the treatment, the MFCs were loaded with a fixed external resistance of 10 Ω and samples of the aqueous solutions in the reactors were collected for COD and sulfate analysis. All of the experiments were conducted at ambient room temperature (ca. 20°C). Given the different designs, chemical molar fluxes into the biofilm on anode electrodes in each reactor were calculated for comparison of chemical utilization rate between the different reactors. Average molar fluxes (mole/cm2/day) of COD and sulfate were calculated as the molar concentration differences over the 14-day treatment divided by anodic surface areas and by the treatment period (i.e., 14 days). Metals and selected elements in the retort water before and after the biological treatments were analyzed using methods described in section 2.7. 2.5. Electric output measurements Electric potential outputs were measured and recorded on a computer using a data acquisition system (USB 6008, National Instrument Inc., USA) equipped with a Labview® software (National Instrument Inc., USA). Electric currents were calculated from the measured electrical potential values and externally loaded resistance. Power densities (mW/m2) of the MFCs were then calculated using the measured potential, calculated currents, and surface areas of the anode electrodes.

Power density curves were developed over a range of external resistance by switching to different pre-installed resistors in a resistor box. 2.6. Coulombic efficiencies Coulombic efficiency was calculated according to Logan et al. (2006). Briefly, the Coulombic efficiency, εC, is defined as the ratio of total Coulombs actually transferred to the anode from the substrate, to maximum possible Coulombs if all removed substrate produced current. The total Coulombs obtained was determined by integrating the current over time in the fed-batch operation, and the Coulombic efficiency for an MFC run in fed-batch mode, εCb, evaluated over a period of time tb, is calculated using Eq. (1) (Cheng et al., 2006). tb

 Cb 

M  I dt 0

Fbv An COD

(1)

where M is the molecular weight of oxygen (32 g/mole); F is Faraday’s constant; b is the number of electrons exchanged per mole of oxygen (4 e-/mole), vAn is the volume of liquid in the anode compartment, and △COD is the change in COD concentration over time tb (14 days). 2.7. Analytical methods Soluble chemical oxygen demand (COD) and sulfate concentrations were measured by a closedreflux method (Standard Method 5220 C) and a turbidimetric method (APHA, 1995), respectively.

1431


Metal concentrations of filtered retort water (0.45 m membrane) were measured using an inductively coupled plasma-optical emission spectrometer (ICP-OES, 710-ES series, Varian Inc., USA). Dissolved oxygen (DO) concentration and the pH of aqueous solutions in the reactors were measured using DO and pH probes with a pH/DO meter (HQ 20, Hach, USA). Particulate material in the aqueous solution of each reactor was collected by filtering the solution and bottom deposits through a membrane filter (0.45 m). The morphology of the particulate material were characterized by scanning electron microscopy (SEM, Hitachi, S-4700) and the chemical composition was analyzed by energy dispersive spectroscopy (EDS, PV7746/58 ME, EDAX Inc., USA) with excitation energy set at 20 kV. The particulate samples were coated with gold under vacuum in a sputter coater (SPI-moduleTM sputter, SPI Supply, USA) prior to the SEM-EDS analysis. In addition, electrode samples were removed from the anode chamber of the MFCs and prepared for SEM-EDS analysis according to Liu et al. (2008). Following an initial wash with sterile water, the electrodes with accumulated biofilm and chemical deposits were then fixed with 2.5% glutaraldehyde in a buffer solution (0.1 mol/L, cacodylate, pH 7.5) at 4°C and then washed with sterile deionized water, followed by stepwise dehydration in a gradient series of water/ethanol solutions (25%, 50%, 75%, 85%, 95%, and 100%). The electrode samples were dried in a desiccator before the SEM-EDS analysis. 3. Results and discussion 3.1. COD utilization and sulfate removal Table 1 lists the initial COD and sulfate concentrations and COD:SO42- ratios used in the experiments. Changes in COD and sulfate during 14 days of treatment by the MFCs are shown in Fig. 2. The initial sulfate concentration in the twochambered MFCs (i.e., MFC-3 and MFC-4) under COD:SO42- ratio 1:1 was raised to examine the effect of high sulfate concentration. Overall, the decreasing trends of both COD and sulfate were similar for the two types of MFCs even though the reactor configurations and electrode surface areas were significantly different. Average molar fluxes of COD and sulfate (i.e., JCOD and Jsulfate) into the biofilm over the 14-day period were calculated (Table 1). For both COD and sulfate, the two-chambered MFCs had molar fluxes 1 to 2 orders of magnitude higher than those of the tubular MFCs. The COD and sulfate molar fluxes for the two-chambered MFCs increased with the COD:SO42- ratio. High sulfate concentrations (e.g., 1,710 mg/L under COD:SO42ratio 1.1) did not result in higher sulfate molar flux than those seen for COD:SO42- ratios 1.5:1 and 2:1, indicating that sulfate removal corresponded more closely with the COD:SO42- ratio than the sulfate concentration alone.

1432

In the tubular MFCs, the averaged sulfate molar flux did not exhibit a strong dependence on the COD:SO42- ratio, indicating that the increase in COD at the higher COD:SO42- did not promote sulfate reduction. It is attributable to oxygen crossover from the cathode to anode due to air bubbling (DO = 0.06 – 2 mg/L in the reactor per measurements). As a result, the anode biofilm may have consisted of aerobic microorganisms on the biofilm surface and SRB embedded underneath the surface. The increased COD was then partially utilized by the aerobic microorganisms with the presence of DO and did not promote sulfate removal significantly. In the two-chambered MFCs, the sulfate flux showed a stronger, positive dependence on the COD:SO42- ratio, suggesting increased sulfate reduction as the organic carbon concentration increased. Although oxygen diffusion from cathode to anode in the two-chambered design has been observed in other studies (Oh et al., 2009), the oxygen crossover from the cathode to anode was likely to be smaller compared to the tubular design. As a result, the increase inorganic carbon contributed mostly to enhanced sulfate reduction, rather than being utilized by aerobic degradation. Rabaey et al. (2006) reported that average sulfate removal from an artificial medium (pH 7.0) in a tubular MFC (0.17 L of liquid volume in anode and 817 ~ 2,720 m2/m3 of granular graphite surface area) was between 2.610-2 and 7.610-2 mmol/m2/day over a 4.2-day period. Zhao et al. (2008) showed that the average sulfate removal from an artificial medium (pH 7.5) in a one-chambered, continuous flow MFC (18 cm3 of liquid volume in the chamber and 1.5 cm2 of anode surface area) with a pure culture of Desulfovibrio desulfuricans and air cathode was 35.3 mmol/m2/day over a 5.8-day period. A comparison of the sulfate removal efficiencies shows that the two-chambered MFCs in this study had larger sulfate molar fluxes using the mixed microbial consortia and the retort water (pH 8.6, ionic strength ~81 mM). 3.2. Sulfur and other elements Analysis of chemical elements in the retort water before and after the MFC treatments showed decreases in the concentrations of total sulfur (S), arsenic, boron, selenium, antimony, strontium and vanadium (data not shown). In particular, sulfur (S) was reduced from 379 mg/L initially to 115 mg/L (MFC-1) and 160 mg/L (MFC-4) after 14 days, which corresponded to sulfate removal percentages of 70% and 58%, respectively. These removal percentages were lower than those of sulfate in the tubular MFCs (92%) and, to a lesser degree, in the two-chambered MFCs (76%) under the same COD:SO42- ratio. Trace amounts of other elements such as iron (Fe), aluminum (Al), nickel (Ni), and thallium (Tl) were also detected in the retort water. The discrepancies between sulfate and sulfur removal


The formation of CuS and/or Cu2S in the MFCs was presumably due to their low solubility products (pKsp ~ 36 and 48). Other elements such as silicon, iron, calcium, and aluminum were also identified. The SEM/EDS data suggested metal sulfides as a possible fate of sulfur in the MFCs. It has been reported that the chemical fate of sulfur in MFCs is largely controlled by anode electropotential (Dutta et al., 2009; Rabaey et al., 2006; Zhao et al., 2008). However, the anode electropotentials were not measured in this study because we focused on the effects of COD:SO42- ratio on sulfate removal. As a result, detailed examination of the mechanisms of the chemical fate of sulfur is not feasible.

(i.e., lower sulfur removal percentages than those of sulfate) may be attributable to formation of other soluble forms of sulfur such as bisulfide, and/or partially oxidized polysulfides and thiosulfate due to presence of limited dissolved oxygen. Similar sulfur balance discrepancy and formation of soluble sulfur forms in MFCs have been reported and suggested by other authors (Rabaey et al., 2006; Zhao et al., 2008). 3.3. SEM/EDS results SEM micrographs and EDS analysis revealed the morphology and chemical composition of particulate matter in the retort water and bottom deposits before and after the biological treatment (Fig. 3). The sample from the original retort water (Fig. 3a) contained mostly NaCl particles (~ 1 m) as indicated by the strong sodium and chloride peaks in the EDS spectrum. The Na and Cl peaks reduced dramatically after the MFC treatments (Figures 3b-3c). Small peaks of sulfur were identified in the particulate matter in the tubular MFC (MFC-1) and the two-chambered MFC (MFC-4), along with calcium, silicon, aluminum, and copper (MFC-4). The SEM micrographs of biofilm and chemical deposits on the anode electrode surfaces show differences in the morphology of the biofilm and chemical composition of the deposits (Fig. 4). Samples of the chemical deposits exhibited copper and sulfur signals, suggesting the formation of CuS or Cu2S on the anode electrodes.

3.4. Electricity generation by the MFCs Fig. 5 presents power curves for electricity generation by the MFCs. The maximum power densities of MFC-1 and MFC-2 (tubular) under COD:SO42- ratio 1.5:1 were 5 mW/m2 and 19 mW/m2, respectively. The maximum power density of MFC-3 and MFC-4 (two-chambered) under COD:SO42- ratio 2:1 were 120 mW/m2 and 106 mW/m2, respectively which were substantially higher than those of the tubular MFCs. In spite of overall higher power produced by the tubular MFCs due to larger electrode surface area (data not shown), the tubular MFCs produced maximum power densities that were around 6 times lower than the two-chambered MFCs.

Table 1. Initial COD and sulfate concentrations, chemical molar flux ratios, and Coulombic efficiencies of the MFCs Initial COD:sulfate ratio 0.5

JCOD (mole/m2/day)

JSulfate (mole/m2/day)

565

Initial sulfate (mg/L) 1131

0.02

0.07

Coulombic Efficiency (%) 2.1

MFC-2

565

1131

0.5

0.01

0.07

2.1

MFC-3

471

950

0.5

1.0

3.4

48.8

MFC-4

471

950

0.5

1.1

4.0

47.1

MFC-1

1156

1168

1.0

0.06

0.07

1.0

MFC-2

1156

1168

1.0

0.09

0.07

0.9

MFC-3

1946

1710

1.1

4.4

4.9

43.3

MFC-4

1946

1710

1.1

4.3

5.1

40.9

MFC-1

1912

1260

1.5

0.09

0.08

0.6

MFC-2

1912

1260

1.5

0.09

0.07

0.6

MFC-3

1692

1130

1.5

4.5

6.6

44.4

MFC-4

1692

1130

1.5

4.6

7.1

44.0

MFC-1

2971

1480

2.0

0.04

0.07

0.4

MFC-2

2971

1480

2.0

0.04

0.07

0.4

MFC-3

2954

1476

2.0

8.7

7.7

25.6

MFC-4

2954

1476

2.0

8.2

7.9

26.1

Reactor

Initial COD (mg/L)

MFC-1

Note: molecular weight of glucose was used for COD molar flux (JCOD) calculation

1433


a)

b)

Fig. 2. Temporal concentration profiles of COD (A) and sulfate (B) in the MFCs under different COD:SO42- ratios over a 14-day period. MFC-1 (△, red), MFC-2 (▽, green), MFC-3 (○, blue), and MFC-4 (□, pink)

Fig. 3. SEM micrographs (5K magnification) and EDS spectra (excitation energy = 20 kV) of the particulate matters in the retort water before the treatment (A) and after treatment using the tubular MFC (B) and two-chambered MFC (C)

The maximum power densities of the twochambered MFCs in this study were higher than a tubular MFC with sulfate oxidation (58 mW/m2) (Rabaey et al., 2006), but were lower than a onechambered, air-breathing cathode and continuous flow MFC with a pure culture of Desulfovibrio desulfuricans (5 W/m2) fed with an artificial wastewater (Zuo et al., 2006). Coulombic efficiencies for all the reactors are listed as a function of COD:SO42- ratio in Table 1. 1434

Overall, the Coulombic efficiencies of the twochambered MFCs were significantly greater than those for the tubular MFCs. This is attributable to greater loss of electrons to DO in the tubular MFCs than in the two-chambered MFCs due to the air bubbling. Electron balance calculation showed that both the Coulombic efficiency and the fractions of electrons released from organic carbon oxidation used for sulfate reduction decreased with COD:SO42- ratio.


Fig. 4. SEM micrographs of biofilm and chemical deposits on the anode electrodes and chemical composition of the accumulated materials. Top panels: SEM images of bacteria (A), chemical deposit (B) and EDS spectrum (C) for the tubular MFC. Bottom panels: SEM image of bacteria (D), chemical deposit (E) and EDS spectrum (F) for the two-chambered MFC

As a result, the fraction of unaccounted electrons (i.e., e- flow to sinks other than electric output and microbial sulfate reduction) increased with the COD:SO42- ratio. 140

Two-chambered MFC-1 Two-chambered MFC-2 Tubular MFC-1 Tubular MFC-2

2

Power density (mW/m )

120 100 80

and electric power output. Low ratios could lead to insufficient removal of sulfate while unnecessarily high ratio could result in dominance of methanogenic bacteria over SRB and low sulfate removal. The possible chemical fates of sulfur and their long-term effects on the anodic microbial activities require more focused studies. This study provides an unique contribution of the effects of COD:SO42- ratio on sulfate removal and power generation using microbial fuel cells. Acknowledgements This work was supported by DOE grant (DOE/RDS/41817M2176). The authors would like to thank James Covell and Mark Thomas for providing the retort water. The authors also thank Charlie Owen at the wastewater treatment plant of the Star City in West Virginia for providing the anaerobic sludge sample.

60 40 20 0 0

200

400

600

800

1000

1200

1400

2

Current density (mA/m )

Fig. 5. Power density curves of the tubular MFCs (MFC-1 and MFC-2) for COD:SO42- ratio 1.5:1 and two-chambered MFCs (MFC-3 and MFC-4) for COD:SO42- ratio 2:1. MFC-1 (△, red), MFC-2 (▽, green), MFC -3 (○, blue), and MFC-4 (□, pink)

4. Conclusions Due to the competing nature of these processes for electron fluxes, organic carbon loading to the MFCs for retort water treatment needs be carefully chosen to achieve optimal sulfate removal

References APHA, (1995), Standard Methods for the Examination of Water and Wastewater, 19th ed. American Public Health Association/American Water Works Association/Water Environment Federation, Washington DC, USA. Bartis J.T., La Tourrette T., Dixon L., Peterson D.J., Cecchine G., (2005), Oil shale development in the United States: Prospects and policy issues, RAND Corporation. Battaglia B.F., Foucher S., Ignatiadis I., Morin D., (2000), Production of hydrogen sulfide by sulfate reducing bacteria in a two-column gas liquid reactor for the purification of metal containing effluents, XXI IMPC, Roma, July 23–27, B12–B17.

1435


Cheng S., Liu H., Logan B.E., (2006), Increased performance of single chamber microbial fuel cells using an improved cathode structure, Electrochemistry Communications, 8, 489–494. Clarke W.P., Ho N.M., Talyor M., Coombs S., Bell P.R., Picaro T., (2005), Evaluation by respirometry of the degradability of retort water using a shale ash and overburden packed column, Environmental Technology, 26, 899–907. Costa M.C., Martins M., Jesus C., Duarte J.C., (2008), Treatment of acid mine drainage by sulfate reducing bacteria using low cost matrices, Water, Air, & Soil Pollution,189, 149–162. Dutta P.K., Keller J., Yuan Z., Rozendal R.A., Rabaey K., (2009), Role of sulfur during acetate oxidation in biological anodes, Environmental Science & Technology, 43, 3839–3845. Dyni J.R., (2006), Geology and resources of some world oil-shale deposits, U. S. Geological Survey Scientific Investigations Report, 42, 2005–5294. Ellis J., Korth J., Peng L., (1995), Treatment of retort waters from Stuart oil shale using high silica zeolites, Fuel, 74, 860–864. Fox J.P., (1981), Elemental Composition of Simulated in situ Oil Shale Retort Water, In: Analysis of Water Associated with Alternative Fuel Production, Hackson L.P., Wright C.C. (Eds.), American Society for testing and materials, ASTM STP 720, West Conshohocken, PA, 101–128. Habermann W.I., Pommer E.H., (1991), Biological fuel cells with sulphide storage capacity, Applied Microbiology and Biotechnology, 35, 128–133. Jeffrey C., Raymond J.K., Paul R.P., (2003), Energy resources and global development, Science, 302, 1528–1531. Liu A., Li H., Liu J., Su Z., (2008), Effects of inoculation strategy and cultivation approach on the performance of microbial fuel cell using marine sediment as biomatrix, Journal of Applied. Microbiology,104, 1163– 1170. Logan B.E., Hamelers B., Rozendal R., Schroder U., Keller J., Freguia S., Aelterman P., Verstraete W., Rabaey

1436

K., (2006), Microbial fuel cells: methodology and technology, Environmental Science & Technology, 40, 5181–5192. Oh S.E., Kim J.R., Joo J.H., Logan B.E., (2009), Effects of applied voltages and dissolved oxygen on sustained power generation by microbial fuel cells, Water Science and Technology, 60, 1311–1317. Petersen K.K., (1981), Oil Shale: the Environmental Challenges II, Proceedings of an International Symposium, August 11-14, 1980, Vail, Colorado, Colorado School of Mines Press, Golden, CO, USA. Rabaey K., Clauwaert P.A., Verstraete W., (2005), Tubular microbial fuel cells for efficient electricity generation, Environmental Science & Technology, 39, 8077–8082. Rabaey K., Sompel K.V.D., Maignien L., Boon N., Aelterman P., Clauwaet P., Schampheaire L.D., Pham H.T., Vermeulen J., Verhaege M., Lens P., Verstraete W., (2006), Microbial fuel cells for sulfide removal, Environmental Science & Technology, 40, 5128–5224. Speight J.G., (2008), Synthetic Fuels Handbook; Properties, Process, and Performance, McGraw-Hill Co., New York. USA. Sebastian J.H., Weber A.S., Jensen J.N., (1996), Sequential chemical/biological oxidation of chlorendic acid, Water Research, 30, 1833–1843. Talve S., Riipulk V., (2001), An inventory analysis of oil shale energy produced on a small thermal power plant, Journal of Cleaner Production, 9, 233–242. Zhang Y., Quan X., Chen S., Zhao Y., Yang F., (2006), Microwave assisted catalytic wet air oxidation of Hacid in aqueous solution under the atmospheric pressure using activated carbon as catalyst, Journal of Hazardous Materials,137, 534–540. Zhao F., Rahunen N., Varcoe J.R., Chandra A., Sossa C.A., Thumser A.E., Slade R.C.T., (2008), Activated carbon cloth as anode for sulfate removal in a microbial fuel cell, Environmental Science & Technology, 42, 4971– 4976. Zuo Y., Maness P.C., Logan B.E., (2006), Electricity production from stem exploded corn stover biomass, Energy & Fuels, 20, 1716–1721.


July 2013, Vol.12, No. 7, 1437-1445



OCCURRENCE AND RISK ASSESSMENT OF ESTROGENS AND ANTIINFLAMMATORIES IN BAIYANGDIAN LAKE, NORTH CHINA Jianghong Shi, Xiaowei Liu, Jinling Cao, Ting Bo, Yingxia Li State Key Laboratory of Water Environment Simulation, School of Environment, Beijing Normal University, Beijing, P.R. China

Abstract Estrogens and anti-inflammatories can pose adverse effects on human health and wildlife. This paper investigated four estrogens (estrone, E1; 17β-estradiol, E2; estriol, E3; 17α-ethynylestradiol, EE2) and three anti-inflammatories (ibuprofen, ketoprofen and naproxen) by solid-phase extraction (SPE) and liquid chromatography tandem mass spectrometry (LC-MS/MS) in Baiyangdian Lake during the dry (April) and wet seasons (September). These compounds were mostly detected at low levels (ng/L). The occurrence and spatial distribution of these compounds varied noticeably with the sampling seasons, particularly antiinflammatories which were present in levels 1–2 orders of magnitude larger in the wet season than in the dry season. The concentrations of these compounds at sampling points S1, S4 and S9 (adjacent to the Fu River), S2 and S8 (affected by human activities) and S3 (receiving the input of more upstream pollutants) were higher than at other sampling points. The overall aquatic environment of Baiyangdian Lake was not at the “high risk” level, but at sites S1, S2 and S3 they were at the “medium risk” level mainly due to the mixture effect of estrogens. At site S2 the hazard quotient (HQ) reached a maximum value of 0.903. This HQ of almost 1 indicates that these estrogens pose a potential environmental risk of induced feminization of male aquatic organisms in Baiyangdian Lake. Key words: anti-inflammatory, Baiyangdian Lake, estrogen, occurrence, risk assessment Received: November 2012; Revised final: May, 2013; Accepted: June 2013

1. Introduction Natural estrogens like estrone (E1), 17βestradiol (E2) and estriol (E3), and the synthetic estrogen 17α-ethynylestradiol (EE2) are excreted in the urine and feces of humans and animals (Preda et al., 2012; Ternes et al., 1999). Anti-inflammatory drugs such as ibuprofen, ketoprofen and naproxen as ‘over the counter’ drugs are widely used as pain relievers and inflammation reducers (Caliman and Gavrilescu, 2009; Preda et al., 2012). These are mainly excreted in the urine and feces of humans (Fent et al., 2006). These compounds have the potential to enter water bodies via inefficiently treated wastewater discharge, runoff from fields applied with manure, or even directly via the introduction of aquaculture. Until recently there have been a growing number of investigations on the 

occurrence and distribution of estrogens and antiinflammatories in surface water. These studies show that these compounds have been widely detected at low levels (in ng/L) in rivers, lakes and other aquatic environments in Switzerland (Tixier et al., 2003), the USA (Boyd et al., 2003), China (Cao et al., 2010; Peng et al., 2008; Wang et al., 2010; ), Italy (Loos et al., 2007; Marchese et al., 2003), the UK (Kasprzyk-Hordern et al., 2008) and South Korea (Kim et al., 2007). Meanwhile, an increasing number of reports (Duong et al., 2010; Fent et al., 2006; Vajda et al., 2008; Ziylan and Ince, 2011) indicate that estrogens and anti-inflammatories may cause potentially adverse effects on the aquatic organisms. Estrogens are well known for their potential endocrine disruptive and reproductive effects on fish (Desbrow et al., 1998; Purdom et al., 1994; Routledge et al.,

Author to whom all correspondence should be addressed: E-mail: [email protected]; Phone: 00861058802846; Fax: 00861058802846

Shi et al./Environmental Engineering and Management Journal 12 (2013), 7, 1437-1445

1998). Hazard and risk assessment studies have been widely conducted for these compounds around the world (Ginebreda et al., 2010; Santos et al., 2007; Schiopu et al., 2012; Wang et al., 2010; Zhao et al., 2011). Use of pharmaceuticals such as antiinflammatories and hormones is increasing in Chinese daily life. In addition, the incomplete removal of pharmaceuticals and estrogens in wastewater treatment plants (WWTPs), and the important ecosystem services provided by rivers and reservoirs like the major freshwater aquaculture production areas and water sources make it is necessary and imperative to investigate the occurrence and distribution of estrogens and antiinflammatories in water bodies. Although there have been several studies on the occurrence and risk assessments for estrogens and pharmaceuticals in the river systems in North China (Cao et al., 2010; Wang et al., 2010), little is known about lakes that provide drinking water or aquaculture water in North China. In order to comprehensively understand the occurrence and risk level of estrogens and anti-inflammatories in the aquatic environments in North China, this study was focused on Baiyangdian Lake, which is the largest natural freshwater lake in the North China Plain providing water resources and aquatic products for the surrounding communities, towns and large cities. 2. Materials and methods 2.1. Study area Baiyangdian Lake, with an area of about 366 km2, is located in the city of Baoding, Hebei Province in North China and is 130 km south of the capital of China, Beijing (Fig. 1). The lake is largely fed by the tributaries of the Daqing River system consisting of the Zhulong River, Tang River, Fu River, Cao River, Pu River and Ping River, and the outflows enter into Bohai Bay via Zhaowangxin Channel. The lake consists of more than 100 small and shallow lakes linked to each other by thousands of ditches, and most parts of the lake are not more than 2 meters in depth. In recent years, due to the construction of upstream reservoirs on the Daqing River and continuous dry years, the shrinking and drying of Baiyangdian Lake has been observed. Compared with the 1950s, the yield of water has been reduced by 10 folds. Due to these changes the tributaries flowing into Baiyangdian Lake have almost dried up, and currently, the Fu River (63 km long and flowing through Baoding City) is the major input into Baiyangdian Lake (Hu et al., 2010). The Fu River is also the discharge channel for sewage and industrial wastewater produced in

1438

Baoding City. Baiyangdian Lake has continuously received domestic sewage and industrial wastewater from Baoding City and nearby towns. In addition, excessive nutrient-rich non-point source pollution from the nearby farmlands is also released into the lake. Baiyangdian Lake is located in the semiarid monsoon climate zone. The average annual precipitation was 557.7 mm in 2008 and precipitation greatly varied through 2008. In the spring (March to May), the average annual precipitation was 53.0 mm, accounting for about 10% of the annual precipitation, while in summer (June to September) the average annual precipitation was 431.0 mm, accounting for about 84% of the annual precipitation. In this study, spring (March to May) was therefore defined as the dry season while summer (June to September) was defined as the wet season. 2.2. Sample collection and site description Two sampling campaigns were carried out during the dry season (April 2008) and the wet season (September 2008). In the dry season, the water storage capacity of Baiyangdian Lake was 67 million m3; while in the wet season, the water storage capacity was 139 million m3, twice the capacity of the dry season. The locations of sampling points are shown in Fig. 1. Surface water samples were collected at three sites (S1 – Wangjiazhai, S4 – Guolikou and S9 – Dazhangzhuang) in Baiyangdian Lake adjacent to wastewater discharge sources via the Fu River in Baoding City (closest to S1), and at three sites (S5 – Duancun, S6 – Datianzhuang and S7 – Caiputai) in the lake adjacent to the Tang River which was a wastewater reservoir (closest to S6). In order to investigate the influence of human activity on the water quality in the lake, S2 – Guangdianzhangzhuang and S8 – Quantou were selected for the sampling points. More importantly, S3 – Zaolinzhuang was selected at the outlet of the lake in order to examine the influence of human activity on the water quality downstream. Amber glass bottles (4 L) were pre-cleaned with tap water, ultrapure water and methanol, and then re-rinsed with sample water three times before sampling. Three parallel samples were collected each time at every sampling point. The pH of the samples was adjusted at the sampling sites to 2.0 with a 4 mol/L HCl. The collected samples were packed in a cool box with ice and transported to the laboratory immediately for further pretreatment and analysis. 2.3. Chemicals and materials The investigated compounds E1, E2, E3, EE2, ibuprofen, ketoprofen and naproxen were purchased from Sigma–Aldrich, USA, all with a purity >95%.

Occurrence and risk assessment of estrogens and anti-inflammatories in Baiyangdian Lake, North China

Fig. 1. Location of sampling points in Baiyangdian Lake

The stock solutions (200 mg/L) containing the seven compounds were prepared in HPLC-grade acetonitrile and were stored at -20 °C. The working solutions were obtained by diluting the stock solutions with acetonitrile/Milli-Q water (1:1, v/v) and were stored at 4 °C. The organic solvents used in this experiment, acetone, acetonitrile, methanol, N-hexane, dichloromethane and ethyl acetate, were HPLC grade (Mallinckrodt Baker Inc., USA), and anhydrous sodium sulfate was analytical reagent grade (Mallinckrodt Baker Inc., USA). The extraction and clean-up of the samples were performed using an Oasis HLB cartridge (200 mg, 6 mL, Waters) and Florisil cartridge (500 mg, 6 mL, Agela), respectively. All glassware that was used in the experiment was soaked in acetone, placed in an ultrasonic bath for 2 h, rinsed with Milli-Q water and baked at 450 °C for 4 h. 2.4. Solid-phase extraction and instrumental analysis Solid-phase extraction (SPE) with an Oasis HLB cartridge was used to separate the seven compounds from the water samples. The HLB cartridge was preconditioned with 5 mL of ethyl acetate, 5 mL of methanol and 10 mL of Milli-Q water. Each water sample (4L) was filtered through glass fiber filters (GF/F, 0.7-μm pore size, Whatman) to remove particulate matter. Then, the filtrate was passed through the HLB cartridge at a flow rate of 3 mL/min under a vacuum. When extraction was completed, the cartridge was washed with 10 mL of Milli-Q water and then dried under vacuum. The cartridge was then eluted with 3×2 mL of ethyl acetate, and the eluent was collected with a 10 mL centrifuge tube. The eluent was then evaporated under a gentle stream of nitrogen and the residue was redissolved with 6 mL of N-hexane/dichloromethane (3:1, v/v). The redissolved sample was passed through a Florisil cartridge (500 mg, 6 mL, Agela), which was filled with a half column of anhydrous sodium sulfate and preconditioned with 5 mL of N-

hexane at a rate of approximately 1 mL/min, and then the Florisil cartridge was washed with 5 mL of Nhexane/ dichloromethane (3:1, v/v) and was eluted with 6 mL of acetone/N-hexane (1:4, v/v). The extract was dried under a gentle nitrogen flow and was reconstituted into 0.2 mL of acetonitrile/water (1:1, v/v). For a liquid chromatography tandem mass spectrometry (LC-MS/MS) analysis, an Agilent 1100 HPLC system (Agilent Technologies, Palo Alto, CA, USA) equipped with a Nova–Pak C18 column (150 mm × 3.9 mm, 4 μm, Waters) was used for the separation of target analytes. A 3200 Q-trap mass spectrometer (Applied biosystems Sciex, USA) was used to quantify the seven target compounds. The four estrogens were separated using the mobile phases A (methanol) and B (water containing 0.2% ammonia, v/v). The gradient for estrogen analysis was performed as follows: from 15% A to 98% A (3 min), 98% A (3 min), and from 98% A to 15% A (4 min). The three anti-inflammatories were separated using the mobile phases A (acetonitrile) and B (water). The gradient for anti-inflammatory analysis was performed as follows: from 15 % A to 90% A (1.5 min), 90% A (2.5 min), and from 90% A to 15% A (5 min). Mass spectrometric analyses were performed in the negative ESI mode. The LC–MS/MS detector was operated in the multiple reaction monitoring (MRM) mode. Detailed information on the performances can be found in the current literature (Zhang et al., 2007; Koh et al., 2007). 2.5. Quality control Quantification of the target compounds was achieved using an external standard method with calibration of seven standard solutions (0.2, 0.5, 1, 2, 5, 10, 20, 50 μg/L) and the correlation coefficients (R2) of the calibration curve were higher than 0.99. Recovery of the entire analytical procedure was

1439


checked through triplicate analysis of water samples spiked with the standard solutions containing the four estrogens and three anti-inflammatories, each at a concentration of 0.5 ng/L. The mean recoveries of spiked samples containing the seven compounds ranged from 73.6– 92.3% and their relative standard deviation (RSD) values were under 20%. The limit of detection (LOD) and the limit of quantitation (LOQ) for instrument were 0.2–0.5 μg/L and 0.5–1 μg/L, respectively, which were determined using the signal-to-noise ratio (SNR). 2.6. Risk assessment In this study, the environmental risk posed by the estrogens and anti-inflammatories on the aquatic ecosystems was assessed using the index of hazard quotients (HQ). This is calculated by dividing the measured concentrations (MC) by the predicted noneffect concentrations (PNEC), as follows Eq. (1).

HQ 

MC PNEC

(1)

PNEC is calculated by dividing the toxicity data (LC50 or lowest chronic no observed effect concentrations (NOEC)) by assessment factors chosen according to the European Technical Guidance Document (EC, 2003). In order to describe the environmental risk more accurately, a more detailed risk ranking criteria was adopted: HQ ≤ 0.1 as “low risk”, 0.1 < HQ ≤ 1 as “medium risk”, and HQ > 1 as “high risk” (Hernando et al., 2006; Wang et al., 2010). 3. Results and discussion The occurrences of estrogens and antiinflammatories in the dry (April) and wet (September) seasons are shown in Fig. 2. Estrogens E1, E2 and EE2 (with the exception of E3) along with the anti-inflammatories ketoprofen, ibuprofen and naproxen were detected at levels of ng/L in Baiyangdian Lake. Results obtained in this study are compared with the levels of these compounds in other water bodies detected by other researchers (Table 1). The occurrence and concentrations of the estrogens and anti-inflammatories demonstrated seasonal variation in the two sampling seasons; in addition, the distribution of these compounds also displayed spatial variation (Fig. 2). 3.1. Seasonal distribution of estrogens Of the four estrogens, E1 was the most frequently detected at all nine sampling points in both dry (April) and wet (September) seasons. E2 was also detected with detection frequencies of 22% and 56% in the dry and wet seasons, respectively, while E3 was not detected. EE2 was only detected in

1440

the dry season with a detection frequency of 56%. With the exception of EE2, the concentrations of these estrogens in the wet season were generally higher than those in the dry season (Fig. 2). For example, concentrations of E1 ranged from 0.02 to 0.19 ng/L, with a mean of 0.04 ng/L in the dry season, and ranged from 0.02 to 0.34 ng/L, with a mean of 0.14 ng/L in the wet season. Concentrations of E2 ranged from ND to 0.03 ng/L, with a mean of 0.02 ng/L in the dry season, while during the wet season they ranged from ND to 0.79 ng/L with a mean of 0.22 ng/L. However, concentrations of EE2 ranged from ND to 0.59 ng/L, with a mean of 0.19 ng/L in the dry season, but it was not detected in the wet season. This finding is contrary to the report of Wang et al. (2012) who found that at most sites concentrations in the dry season were generally higher than those in the wet season. The slightly higher concentrations of E1 and E2 in the wet season is possible due to the fact that cleavage of the inactive polar conjugates of E1 and E2 occurred more easily in the wet season because of the higher water temperature (23.6–26.0 °C) than during the dry season (16.6–19.5 °C). In contrast, EE2 was detected only during the dry season. This may be caused by a bigger dilution effect playing a dominant role in the wet season. Table 1 shows that concentrations of E1 in Baiyangdian Lake were significantly lower than those reported in the Pearl River in South China (Peng et al., 2008), Yellow River in North China (Wang et al., 2012), surface water in Korea (Kim et al., 2007), and rivers in the U.S. (Kolpin et al., 2002). However, concentrations of E1 in Baiyangdian Lake were comparable to those reported in the Lake Maggiore in Northern Italy (Loos et al., 2007) and Llobregat River Basin of Spain (Kusteret al., 2008). Concentrations of E2 were similar to the levels reported of all other noted regions, with the exception of the U.S. Concentrations of EE2 were also comparable to those report in the Pearl River in South China (Peng et al., 2008), while they were lower than those in the Venice lagoon in Italy (Pojana G et al., 2007) and the rivers in the U.S. (Kolpin et al., 2002). 3.2. Seasonal distribution of anti-inflammatories The three anti-inflammatories were detected at the low levels (ng/L) and their concentrations clearly varied with the change in seasons (Fig. 2). In both sampling seasons ibuprofen and ketoprofen were detected with detection frequencies of 22% and 100% in the dry season (April), and 44% and 44% in the wet season (September), respectively. Ibuprofen was detected in both sampling seasons; this may be due to its status as one of the four most often-used anti-inflammatories. Ketoprofen was widely detected, but it was in contrast to the report of Wang et al. (2010) in which ketoprofen was not detected. In the dry season,

Occurrence and risk assessment of estrogens and anti-inflammatories in Baiyangdian Lake, North China

mainly in conjugation with glucuronic acid (Ziylan and Ince, 2011); with the increase of temperature, cleavage of the conjugates of ketoprofen occur more easily causing higher concentrations during the wet season. Therefore, it can be concluded that during the wet season the concentration decreases of the three anti-inflammatories due to the degradation under higher temperatures and the dilution effect caused by the large runoff were slight compared with the cleavage of the conjugates and non-point source pollution. The levels of three anti-inflammatories in Baiyangdian Lake are relatively low in comparison with in other water bodies (Table 1). These concentrations are even 4 orders of magnitude lower than those in the Llobregat River (NE Spain) (Ginebreda et al., 2009). Nevertheless, the concentration range of ibuprofen in Baiyangdian Lake was similar to those reported in Lake Maggiore in Northern Italy (Loos et al., 2007) and in the Mississippi River in the U.S. (Boyd et al., 2004); the concentration ranges of naproxen and ketoprofen were comparable to those in the Vico Lake and Bracciano Lake in Italy (Marchese et al., 2003).

ibuprofen and ketoprofen had concentrations ranging from ND to 0.04 ng/L (for ibuprofen) and 0.04 to 0.59 ng/L (for ketoprofen), and mean concentrations of 0.02 and 0.10 ng/L, respectively. In the wet season, naproxen was the most frequently detected anti-inflammatory, at a detection frequency of 89%, although all three anti-inflammatories were detected. The concentrations of ibuprofen, naproxen and ketoprofen ranged from ND to 2.19, ND to 2.95, and ND to 30.8 ng/L, respectively. These compounds had mean concentrations of 1.40, 1.70 and 9.98 ng/L. Compared with the dry season, the detection concentrations in the wet season were 1–2 orders of magnitude higher than those in the dry season. During the wet season, with higher temperatures the cleavage of conjugates of the three antiinflammatories occurred more easily. In addition, the non-point source pollution resulting from large amount of runoff might be responsible for the higher concentrations in the wet season. Moreover, the concentrations of ketoprofen in the wet season (a mean concentration of 9.98 ng/L) were far higher than those in the dry season (a mean concentration of 0.10 ng/L). This can also be explained by the fact that in this environment ketoprofen is metabolized

2.0

April September

0.3

E1

Ibuprofen

1.5 1.0

0.2

0.05

0.1

Concentration (ng/L)

0.0

S1

S2

S3

S4

S5

S6

S7

S8

S9

0.00

S1

S2

S3

S4

S5

E2

S8

S9

Naproxen

0.7

2 1

0.1

S1

S2

S3

S4

S5

S6

0.6

S7

S8

S9

0

4

0.2

2

S1

S2

S3

S4

S5

S6

S7

S1

S2

S3

S4

S5

30 20

EE2

0.4

0.0

S7

3

0.8

0.0

S6

S8

S9

0

S6

S7

S8

S9

Ketopronfen

S1

S2

S3

S4

S5

S6

S7

S8

S9

Sampling points Fig. 2. Seasonal and spatial distribution of estrogens (E1, E2 and EE2) and anti-inflammatories (ketoprofen, ibuprofen and naproxen) in Baiyangdian Lake during the dry (April) and wet (September) seasons. The error bars indicate the standard deviations of the measured concentrations (n = 3)

1441


Table 1. Comparison of concentrations (ng/L) of the seven compounds in the Baiyangdian Lake and other water bodies in the world Baiyangdian Lake, North China E1

0.02-0.34

Pearl River, South China ND-65

Yellow River, North China ND-15.6

Rivers, Spain

Surface water, Korea

Rivers and lakes, US

River and Lakes, Italy

0.75-1.68

1.7-5.0

NQ-112

ND-0.4

E2

ND-0.79

ND-2

ND-2.3

NA

ND