Natural Arsenic in Groundwater: Occurrence, Remediation and Management Proceedings of the Pre-Congress Workshop 'Natural Arsenic in Groundwater (BWO 06), ... Congress, Florence, Italy, 18-19 August 2004

NATURAL ARSENIC IN GROUNDWATER: OCCURRENCE, REMEDIATION AND MANAGEMENT

PROCEEDINGS OF THE PRE-CONGRESS WORKSHOP “NATURAL ARSENIC IN GROUNDWATER (BWO 06)”, 32nd INTERNATIONAL GEOLOGICAL CONGRESS, FLORENCE, ITALY, 18–19 AUGUST 2004

Natural Arsenic in Groundwater: Occurrence, Remediation and Management

Edited by

Jochen Bundschuh International Technical Co-operation Programme CIM (GTZ/BA), Frankfurt, Germany – Instituto Costarricense de Electricidad ICE, San José, Costa Rica

Prosun Bhattacharya Groundwater Arsenic Research Group, Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden

D. Chandrasekharam Department of Earth Sciences, Indian Institute of Technology, Bombay, India

A.A. BALKEMA PUBLISHERS

LEIDEN / LONDON / NEW YORK / PHILADELPHIA / SINGAPORE

This edition published in the Taylor & Francis e-Library, 2005. “To purchase your own copy of this or any of Taylor & Francis or Routledge’s collection of thousands of eBooks please go to www.eBookstore.tandf.co.uk.”

Copyright © 2005 Taylor & Francis Group plc, London, UK All rights reserved. No part of this publication or the information contained herein may be reproduced, stored in a retrieval system,or transmitted in any form or by any means, electronic, mechanical, by photocopying, recording or otherwise, without written prior permission from the publisher. Although all care is taken to ensure the integrity and quality of this publication and the information herein, no responsibility is assumed by the publishers nor the author for any damage to property or persons as a result of operation or use of this publication and/or the information contained herein. Published by: A.A. Balkema Publishers, Leiden, The Netherlands, a member of Taylor & Francis Group plc www.balkema.nl and www.tandf.co.uk ISBN 0-203-97082-9 Master e-book ISBN

ISBN 04 1536 700 X (Print Edition)

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Table of Contents

Preface

IX

List of Contributors

XI

Section 1: Arsenic occurrence and genesis in sedimentary and hard-rock aquifers Arsenic in groundwater of the Bengal Delta Plain: geochemical evidences for small scale redox zonation in the aquifer F. Wagner, Z.A. Berner & D. Stüben

3

Genesis of arsenic contamination of groundwater in alluvial Gangetic aquifer in India S.K. Acharyya & B.A. Shah

17

Arsenic pollution in groundwater of West Bengal, India: Where we stand? D. Chandrasekharam

25

Mineralogical characteristics of the Meghna floodplain sediments and arsenic enrichment in groundwater A.M. Sikder, M.H. Khan, M.A. Hasan & K.M. Ahmed

31

Naturally occurring arsenic in groundwater of Terai region in Nepal and mitigation options N. Tandukar, P. Bhattacharya, G. Jacks & A.A. Valero

41

High arsenic concentrations in mining waters at Kanˇk, Czech Republic A. Koprˇiva, J. Zeman & O. Sracek

49

Natural arsenic in the groundwater of the alluvial aquifers of Santiago del Estero Province, Argentina P. Bhattacharya, M. Claesson, J. Fagerberg, J. Bundschuh, A.R. Storniolo, R.A. Martin, J.M. Thir & O. Sracek

57

Arsenic source and fate at a village drinking water supply in Mexico and its relationship to sewage contamination J.M. Cole, M.C. Ryan, S. Smith & D. Bethune

67

Arsenic contamination of the Salamanca aquifer system in Mexico: a risk analysis R. Rodriguez, M.A. Armienta & J.A. Mejia Gómez

77

Arsenic pollution in aquifers located within limestone areas of Ogun State, Nigeria A.M. Gbadebo

85

Section 2: Environmental health assessment-arsenic in the food chain Arsenic in groundwater and contamination of the food chain: Bangladesh scenario S.M.I. Huq & R. Naidu Arsenic contamination in groundwater in Nepal: a new perspective and more health threat in South Asia S.R. Kanel, H. Choi, K.W. Kim & S.H. Moon V

95 103

Estimating previous exposure to arsenic for populations living in parts of Hungary, Romania and Slovakia R.L. Hough, G.S. Leonardi & T. Fletcher

109

Arsenic bioaccumulation in a green algae and its subsequent recycling in soils of Bangladesh S.M.I. Huq, A. Bulbul, M.S. Choudhury, S. Alam & S. Kawai

119

Environmental behavior of arsenic in a mining zone: Zimapán, Mexico M.A. Armienta, R. Rodríguez, O. Cruz, A. Aguayo, N. Ceniceros, G. Villaseñor, L.K. Ongley & H. Mango

125

Section 3: Arsenic biogeochemistry in groundwater Natural enrichment of arsenic in groundwaters of Brahmanbaria district, Bangladesh: geochemistry, speciation modeling and multivariate statistics O. Sracek, P. Bhattacharya, M. von Brömssen, G. Jacks & K.M. Ahmed

133

Microbial processes and arsenic mobilization in mine tailings and shallow aquifers J. Routh & A. Saraswathy

145

Geochemistry and geomicrobiology of arsenic in Holocene alluvial aquifers, USA J.A. Saunders, M.K. Lee & S. Mohammad

155

Arsenic contamination in drinking water of tube wells in Bangladesh: statistical analysis and associated factors M.A. Hossain, M. Amirul Islam, M.O. Gani & M.A. Karim

163

The impact of low dissolved oxygen in recharge water on arsenic pollution in groundwater of Bangladesh Md. N. Islam & R.D. Bob von Bernuth

173

Section 4: Remediation of arsenic-rich groundwaters Technologies for arsenic removal from potable water W. Driehaus

189

Natural enrichment of arsenic in a minerotrophic peatland (Gola di Lago, Canton Ticino, Switzerland), and implications for the treatment of contaminated waters Z.I. González-Acevedo, M. Krachler, A.K. Cheburkin & W. Shotyk

205

A comparative study for the removal of As(III) and As(V) by activated alumina T.S. Singh & K.K. Pant

211

Comparing the arsenic sorption capacity of Bauxsol™ and its derivatives with other sorbents H. Genç-Fuhrman, D. McConchie & O. Schuiling

223

Optimization of the removal of arsenic from groundwater using ion exchange C.N. Mulligan, A.K.M. Saiduzzaman & J. Hadjinicolaou

237

Sorption of arsenic on sorghum biomass: a case study N. Haque, G. Morrison, G. Perrusquía, I. Cano-Aguilera, A.F. Aguilera-Alvarado & M. Gutiérrez-Valtierra

247

Removal and recovery of arsenic from aqueous solutions by sorghum biomass Z.I. González-Acevedo, I. Cano-Aguilera & A.F. Aguilera-Alvarado

255

VI

Optimisation of iron removal units to include arsenic removal A.K. Sharma, J.C. Tjell & H. Mosbæk

263

A simple and environmentally safe process for arsenic remediation – laboratory and field evaluation K. Misra, M.T. Companywala, S. Sharma, A. Srivastava & P.C. Deb

273

Section 5: Management of arsenic-rich groundwaters Management of the groundwater arsenic disaster in Bangladesh K.M. Ahmed

283

Strengthening water examination system in Bangladesh H. Jigami

297

Implementation of safe drinking water supplies in Bangladesh C.F. Rammelt & J. Boes

307

Sustainable safe water options in Bangladesh: experiences from the Arsenic Project at Matlab (AsMat) Md. Jakariya, Mizanur Rahman, A.M.R. Chowdhury, Mahfuzar Rahman, Md. Yunus, A. Bhiuya, M.A. Wahed, P. Bhattacharya, G. Jacks, M. Vahter & L.-Å. Persson

319

Prerequisite studies for numerical flow modeling to locate safe drinking water wells in the zone of arsenic polluted groundwater in the Yamuna sub-basin, West Bengal, India S. Mukherjee

331

Author index

339

VII

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Preface

Groundwater is an important resource that serves as a backbone of human development. In several regions, mostly in developing countries, groundwater from sedimentary and hard rock aquifers used for drinking is naturally contaminated with arsenic. In different countries in Asia such as India, Bangladesh, Cambodia, China, Nepal, Pakistan, Taiwan, Thailand and Vietnam, the situation of arsenic toxicity is alarming and severe health problems are reported amongst the inhabitants relying on groundwater as sources of water for drinking purposes. Arsenic occurrences in groundwater in Bengal Delta Plain of West Bengal, India and Bangladesh is one of the largest environmental health disasters of the present century, where over 50 million people are at risk of cancer and other arsenic related diseases. In these same countries, land and agricultural sustainability is threatened by the use of arsenic contaminated irrigation water. In several Middle- and South-American countries, for example in Argentina, Brazil, Chile and Mexico, high arsenic is reported in natural waters. In Argentina, at least 1.2 million people are affected. Elevated levels of natural arsenic in groundwater due to geogenic sources is therefore an issue of primary environmental concern, which limits the use of these resources for drinking or other purposes, and hinders the socio-economic growth. Hence there is a need to improve our understanding on the genesis of high arsenic groundwaters from the various aquifers in order to develop strategies to save millions from this calamity. This publication comprises articles concerning the occurrence of arsenic in groundwater, its mobility constraints, water-sediment interactions, various other related themes concerning aqueous geochemistry of arsenic in sedimentary and hard rock aquifers and other natural environmental systems around the world, assessment of environmental health risks and impacts, and the arsenic removal technologies. These articles are based on the papers presented during the Pre-Congress Workshop Natural Arsenic in Groundwater (BWO 06), which was held on 18–19 August 2004, as an event of the 32nd International Geological Congress in Florence, Italy between August 20–28, 2004. The book is divided into five sections, covering the following themes: 1. Arsenic occurrence in sedimentary and hard-rock aquifers: Case studies from Argentina, India, Bangladesh, Nepal, Nigeria, Czech Republic and Mexico are presented in 10 Chapters. 2. Environmental health assessment – Arsenic in the food chain: Fate of arsenic in soil, water and crops - extent, sources, characterization and background levels, implications to human health, south Asian experience, toxicological impacts and environmental health assessment of arsenic poisoning are discussed in 5 Chapters. 3. Arsenic biogeochemistry in groundwater: This section deals with the speciation and mobility controls of arsenic, mechanisms of redox transformations through biotic and abiotic processes, influence of microbiota and the microbial reactions and their implication for arsenic mobilization, speciation modeling, adsorption/desorption processes, thermodynamic constraints on arsenic solubility and statistical analyses of hydrogeochemical parameters, presented in 5 Chapters. 4. Remediation of arsenic rich groundwaters: The removal of arsenic from groundwater is an important global issue. Several millions of people are drinking water with elevated levels of arsenic compared to the drinking water standards recommended by World Health Organization (WHO). Various of treatment technologies are available for removal of arsenic worldwide, which range from sophisticated ion exchange and reverse osmosis, simple conventional coagulation-flocculation technique and small scale filters. Incidences of elevated arsenic concentrations in groundwater within the developing countries with poor infrastructure demand technologies that are effective and affordable for the provision of safe drinking water supply to the affected population. This section comprises 9 different Chapters, which discuss various low IX

cost and environment friendly techniques of arsenic removal that are suitable for providing safe drinking water to the affected population in various parts of the world. 5. Management of arsenic rich groundwaters: The natural background concentration of arsenic in soils is an important factor to assess the environmental quality and strategies for subsequent remediation. Remediation of arsenic contaminated groundwater systems are much more complicated and often involves designing economically feasible and effective techniques that are site-specific. Effective strategies for groundwater management are needed to circumvent the environmental health disasters, which are focused in the last 5 Chapters. The book forms a base for discussion and exchange of scientific ideas to identify future targets for research needed to improve the understanding of the mobility of arsenic in the aquifers and their impacts. The book is designed to: (i) create interest within the countries which are affected by arseniferous aquifers; (ii) to update the current status of knowledge on the dynamics of natural arsenic from the aquifers through groundwater to food chain for professionals involved in the topic; (iii) an important worldwide issue on improved and efficient techniques for arsenic removal in regions with elevated arsenic levels in groundwater; and (iv) bring awareness, among administrators, policy makers and company executives, on the problem and to improve the international cooperation on that topic. We would like to thank the organizers of the 32nd International Geological Congress, who have given us the opportunity to organize the Pre-Congress Workshop ‘Natural Arsenic in Groundwater’ (BWO 06) and the infrastructural facility to stage the event. We would like to thank our colleagues Kazi Matin Ahmed, Alan Welch, Andreas Mende, Klaus-Peter Seiler, Doris Stüben, Ondra Sracek, Gunnar Jacks, Maria Aurora Armienta, Richard Johnston, Mahfuzar Rahman, Jim Saunders, S.K. Acharyya, William Burgess and D.K. Guha for their efforts with the timely review of the manuscripts of the Chapters in this book. We wish to express our sincere thanks to them, who contributed significantly to maintain the high quality of the papers in this volume. We would like to thank the Swedish International Development Agency (Sida-SAREC), Swedish Research Council (VR), Swedish Research Council for Environment, Agricultural Sciences and Spatial Planning (Formas) and the GeoHost programme of the International Geological Congress for support to the participants in this Workshop. We are especially grateful to the organisational support and cooperation provided by the Integrated Expert Programme of CIM (GTZ/BA), Frankfurt, Germany, the Instituto Costarricense de Electricidad (ICE), San José, Costa Rica, the Universidad Nacional de Santiago del Estero (UNSE), Argentina, the Royal Institute of Technology (KTH), Stockholm, Sweden and the Indian Institute of Technology Bombay, India. We thank the GEH Wasserchemie GmbH & Co. KG, Osnabrück, Germany, for their contribution to the printing costs of this book. Jochen Bundschuh Prosun Bhattacharya D. Chandrasekharam (Editors)

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

List of Contributors

S.K. Acharyya Department of Geological Sciences, Jadavpur University, Kolkata, India A. Aguayo Instituto de Geofísica, UNAM, México D.F., Mexico A.F. Aguilera-Alvarado Facultad de Química, Universidad de Guanajuato, Guanajuato, Gto., Mexico K.M. Ahmed Department of Geology, University of Dhaka, Dhaka, Bangladesh S. Alam Department of Agro-Bioscience, Iwate University, Morioka, Japan M.A. Armienta Instituto de Geofisica, Universidad Nacional Autónoma de México (UNAM), C.U., Del Coyoacan, México City, México Z.A. Berner Institute of Mineralogy and Geochemistry, Universität Karlsruhe (TH), Karlsruhe, Germany D. Bethune Department of Geology and Geophysics, University of Calgary, Calgary, Alberta, Canada A. Bhiuya Center for Health and Population Research, International Centre for Diarrhoeal Disease Research, Bangladesh (ICDDR,B), Dhaka, Bangladesh P. Bhattacharya Groundwater Arsenic Research Group, Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden J. Boes Faculty of Technology, Policy and Management, Delft University of Technology, Delft, Netherlands M. von Brömssen Groundwater Arsenic Research Group, Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden R.D. von Bernuth Department of Biosystems Engineering, Michigan State University, East Lansing, Michigan, USA A. Bulbul Department of Soil, Water & Environment, University of Dhaka, Dhaka, Bangladesh XI

J. Bundschuh International Technical Co-operation Programme CIM(GTZ/BA), Frankfurt, Germany – Instituto Costarricense de Electricidad ICE, PySA, Costa Rica; Facultad de Ciencias Exactas y Tecnologias, Universidad Nacional de Santiago del Estero (UNSE), Santiago del Estero, Argentina I. Cano-Aguilera Facultad de Química, Universidad de Guanajuato, Guanajuato, Gto., Mexico N. Ceniceros Instituto de Geofísica, UNAM, México D.F., Mexico D. Chandrasekharam Department of Earth Sciences, Indian Institute of Technology Bombay, Mumbai, India A.K. Cheburkin Institute of Environmental Geochemistry, Heidelberg University, Heidelberg, Germany H. Choi Gwangju Institute of Science and Technology, Puk-gu, Gwangju, Republic of Korea M.S. Choudhury Department of Soil, Water & Environment, University of Dhaka, Dhaka, Bangladesh A.M.R. Chowdhury Research and Evaluation Division, BRAC, Dhaka, Bangladesh M. Claesson Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden J.M. Cole Department of Geology and Geophysics, University of Calgary, Calgary, Alberta, Canada M.T. Companywala Naval Materials Research Laboratory (NMRL), DRDO, Ministry of Defence, Addl. Ambernath, India O. Cruz Instituto de Geofísica, UNAM, México D.F., Mexico P.C. Deb Naval Materials Research Laboratory (NMRL), DRDO, Ministry of Defence, Addl. Ambernath, India W. Driehaus GEH Wasserchemie GmbH & Co. KG, Osnabrueck, Germany J. Fagerberg Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden T. Fletcher London School of Hygiene & Tropical Medicine, University of London, London, UK M.O. Gani Graduate Student, Faculty of Agricultural Engineering & Technology, Bangladesh Agricultural University, Mymensingh, Bangladesh XII

A.M. Gbadebo Department of Environmental Management and Toxicology, College of Environmental Management Resources, University of Agriculture, Abeokuta, Ogun State, Nigeria H. Genç-Fuhrman Environment & Resources, Technical University of Denmark (DTU), Lyngby, Denmark Z.I. González -Acevedo Facultad de Química, Universidad de Guanajuato, Guanajuato, Gto., México Z.I. González -Acevedo Institute of Environmental Geochemistry, Heidelberg University, Heidelberg, Germany M. Gutiérrez-Valtierra Facultad de Química, Universidad de Guanajuato, Guanajuato, Gto., Mexico J. Hadjinicolaou Department of Building, Civil and Environmental Engineering, Concordia University, Montreal, Quebec, Canada M.A. Hasan Department of Geology, University of Dhaka, Dhaka, Bangladesh N. Haque Water Environment Transport, Chalmers University of Technology, Göteborg, Sweden M.A. Hossain Department of Farm Structure, Faculty of Agricultural Engineering & Technology, Bangladesh Agricultural University, Mymensingh, Bangladesh R.L. Hough London School of Hygiene & Tropical Medicine, University of London, London, UK S.M.I. Huq Department of Soil, Water & Environment, University of Dhaka, Dhaka, Bangladesh Md. N. Islam Department of Civil Engineering, University of Toronto, Toronto, Canada G. Jacks Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden Md. Jakariya Research and Evaluation Division, BRAC, Dhaka, Bangladesh H. Jigami Japan International Cooperation Agency (JICA) Expert, Department of Public Health Engineering, Dhaka, Bangladesh S.R. Kanel Gwangju Institute of Science and Technology, Puk-gu, Gwangju, Republic of Korea M.A. Karim Graduate Student, Faculty of Agricultural Engineering & Technology, Bangladesh Agricultural University, Mymensingh, Bangladesh S. Kawai Department of Agro-Bioscience, Iwate University, Morioka, Japan XIII

M.H. Khan Arsenic Research Group [BD], Banani, Dhaka, Bangladesh K.W. Kim Gwangju Institute of Science and Technology, Puk-gu, Gwangju, Republic of Korea A. Koprˇiva Institute of Geological Sciences, Faculty of Science, Masaryk University, Brno, Czech Republic M. Krachler Institute of Environmental Geochemistry, Heidelberg University, Heidelberg, Germany M.K. Lee Department of Geology and Geography, Auburn University, Auburn, USA G.S. Leonardi London School of Hygiene & Tropical Medicine, University of London, London, UK H. Mango Dept. of Natural Sciences, Castleton State College, Castleton, USA R. A. Martin Facultad de Ciencias Exactas y Tecnologias, Universidad Nacional de Santiago del Estero (UNSE), Santiago del Estero, Argentina D. McConchie Southern Cross University, Centre for Coastal Management, East Lismore, Australia J.A. Mejia Gómez Consejo Tecnico del Agua, COTAS, Irapuato-Valle de Santiago, Gto., Mexico K. Misra Naval Materials Research Laboratory (NMRL), DRDO, Ministry of Defence, Addl. Ambernath, India S. Mohammad Department of Geology and Geography, Auburn University, Auburn, USA S.H. Moon Gwangju Institute of Science and Technology, Puk-gu, Gwangju, Republic of Korea G. Morrison Water Environment Transport, Chalmers University of Technology, Göteborg, Sweden H. Mosbæk Environment & Resources, Technical University of Denmark (DTU), Lyngby, Denmark C.N. Mulligan Department of Building, Civil and Environmental Engineering, Concordia University, Montreal, Quebec, Canada S. Mukherjee Department of Earth Sciences, Indian Institute of Technology Roorkee, Roorkee, Uttaranchal, India L.K. Ongley Oak Hill High School, Sabatus, USA XIV

R. Naidu Centre for Environmental Risk Assessment and Remediation (CERAR), University of South Australia, Mawson Lake Campus, Adelaide, Australia K.K. Pant Department of Chemical Engineering, Indian Institute of Technology, New Delhi, India G. Perrusquía Water Environment Transport, Chalmers University of Technology, Göteborg, Sweden L.Å. Persson Department of Women’s and Children’s Health, Uppsala University, Uppsala, Sweden Mahfuzar Rahman Center for Health and Population Research, International Centre for Diarrhoeal Disease Research, Bangladesh (ICDDR,B), Dhaka, Bangladesh Mizanur Rahman Research and Evaluation Division, BRAC , Dhaka, Bangladesh C.F. Rammelt Faculty of Technology, Policy and Management, Delft University of Technology, Delft, Netherlands R. Rodriguez Instituto de Geofisica, Universidad Nacional Autónoma de México (UNAM), C.U., Del Coyoacan, México City, México J. Routh Department of Geology and Geochemistry, Stockholm University, Stockholm, Sweden M.C. Ryan Department of Geology and Geophysics, University of Calgary, Calgary, Alberta, Canada A.K.M. Saiduzzaman Department of Building, Civil and Environmental Engineering, Concordia University, Montreal, Quebec, Canada A. Saraswathy Department of Biology, West Virginia State College, West Virginia, USA J.A. Saunders Department of Geology and Geography, Auburn University, Auburn, USA O. Schuiling Institute of Earth Sciences, Utrecht, The Netherlands B.A. Shah Department of Geological Sciences, Jadavpur University, Kolkata, India A.K. Sharma Environment & Resources, Technical University of Denmark (DTU), Lyngby, Denmark S. Sharma Naval Materials Research Laboratory (NMRL), DRDO, Ministry of Defence, Addl. Ambernath, India W. Shotyk Institute of Environmental Geochemistry, Heidelberg University, Heidelberg, Germany XV

A.M. Sikder Arsenic Research Group [BD], 306 Iqbal Center, Banani, Dhaka, Bangladesh T.S. Singh Department of Chemical Engineering, Indian Institute of Technology, New Delhi, India S. Smith Caminamos Juntos para Salud y Desarrollo, A.C. Cuernavaca, Morelos, México O. Sracek Institute of Geological Sciences, Faculty of Science, Masaryk University, Brno, Czech Republic A. Srivastava Naval Materials Research Laboratory (NMRL), DRDO, Ministry of Defence, Addl. Ambernath, India A.R. Storniolo Facultad de Ciencias Exactas y Tecnologias, Universidad Nacional de Santiago del Estero (UNSE), Santiago del Estero, Argentina D. Stüben Institute of Mineralogy and Geochemistry, Universität Karlsruhe (TH), Karlsruhe, Germany N. Tandukar Department of Water Supply & Sewerage (DWSS), Kathmandu, Nepal J.M. Thir Facultad de Ciencias Exactas y Tecnologias, Universidad Nacional de Santiago del Estero (UNSE), Santiago del Estero, Argentina J.C. Tjell Environment & Resources, Technical University of Denmark (DTU), Lyngby, Denmark M. Vahter Division of Metals and Health, Karolinska Institutet, Stockholm, Sweden A.A. Valero Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), SE-10044 Stockholm, Sweden G. Villaseñor Instituto de Geología, UNAM, México D.F. F. Wagner Institute of Mineralogy and Geochemistry, Universität Karlsruhe (TH), Karlsruhe, Germany M.A. Wahed Center for Health and Population Research, International Centre for Diarrhoeal Disease Research, Bangladesh (ICDDR,B), Dhaka, Bangladesh Md. Yunus Center for Health and Population Research, International Centre for Diarrhoeal Disease Research, Bangladesh (ICDDR,B), Dhaka, Bangladesh J. Zeman Institute of Geological Sciences, Faculty of Science, Masaryk University, Brno, Czech Republic

XVI

Section 1: Arsenic occurrence and genesis in sedimentary and hard-rock aquifers

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Arsenic in groundwater of the Bengal Delta Plain: geochemical evidences for small scale redox zonation in the aquifer F. Wagner, Z.A. Berner & D. Stüben Institute of Mineralogy and Geochemistry, Universität Karlsruhe (TH), Karlsruhe, Germany

ABSTRACT: The study focuses on characterisation of the hydrogeochemical environment of high As concentrations in groundwater and on the availability and mineralogical speciation of As in the aquifer sediments of an area of about 35 km2 (Malda District, West Bengal, India), where hot-spots with As enriched groundwater occur in close vicinity of low As zones. Evidence has been found for the occurrence of distinct redox zones with high As concentrations, in which conditions for selective reduction of Mn, Fe or SO4 occur. This implies that under specific local conditions Mn- and Fe-oxihydroxides represent important As-bearing phases in the aquifer. The comparison of the results of two field campaigns carried out in 2002 and 12 month later indicates a shift towards slightly stronger reducing conditions, along with local increases of arsenic concentrations in groundwater. High As concentrations in the shallow groundwater apparently are connected to abandoned river channels. In such oxbow channels fine grained sediments rich in Corg are typically deposited, which may accelerate the development of a reducing environment due to the microbial mineralization of organic matter. The percolation of sewage into the aquifer as an additional anthropogenic source of nutrients may be suggested, but its importance for the development of reducing conditions in the groundwater is hard to estimate, considering the regional scale of the As calamity. Statistical evaluation of ␮-XRFA data suggests Fe bearing silicates (such as chlorite and biotite) as the main As carrier mineral phases in the aquifer sediments. Nevertheless, the redox-sensitive mobilization of As from disperse distributed Fe- and Mn-oxihydroxides is likely to be considered as the main driver for the enrichment of As in groundwater.

1

INTRODUCTION

The groundwater of the Bengal Delta Plain, the main source of drinking water for many millions of people and intensively used for irrigation of large agricultural areas, has As concentrations which exceed the 10 ␮g/L level recommended by WHO by two order of magnitude. Several studies were conducted to elucidate the cause of this calamity and to work out and implement measures in order to mitigate its impact on population since the regional scale of the problem became manifest in the early 1990s. However, despite the intensive effort of different scientific, economic and governmental institutions, on both national and international scale, the problem is far from solved. Though the geogenic source of the As in the groundwater is generally accepted, the primary source and mechanism of release of As from the aquifer sediments into the groundwater is still not well understood. The role of human activities in leading to the increase of As concentrations – like e.g., the intensive exploitation of groundwater for irrigation – is still a matter of intensive debate. Earlier investigators considered that As is released when pyrite is oxidized due to the aeration of the aquifer sediments as a consequence of intensive groundwater exploitation. However, this assumption as such is hardly sustainable, because no correlation has been found between As occurrence in groundwater and groundwater extraction. Furthermore, arsenic may rapidly be adsorbed on the surface of the Fe-oxides formed by pyrite weathering. Lately the reduction of arsenic-bearing iron oxides is becoming an increasingly popular model to explain the high As 3

contents in groundwaters of the BDP (Nickson et al. 1998, BGS & DPHE 2001, Stüben et al. 2002). Nevertheless, a consistent model is needed to incorporate the multitude of accumulated observations. It needs to be clarified what is the primary source, or more exactly the mineral speciation and availability/mobility of As in the aquifer sediments, what determine the decrease of the redox state and what is the role of organic matter (e.g. McArthur et al. 2001, Ravenscroft et al. 2001). If arsenic is released by the reduction of Fe and Mn-oxihydroxides, how can be explained the often missing correlation among these elements and what is the reason behind the irregular “patchy” distribution of the As-rich zones. More data are needed also on the evolution of the As concentration in groundwater with time. The present study focuses on some of these aspects, through a detailed geochemical study on both groundwater and aquifer sediments in an area with As enriched hot spots in close proximity to low As zones. Additionally, further study on the impact of irrigation with arsenic rich groundwater on rice, wheat and agricultural soil in the same area was conducted (Norra et al. submitted).

2

METHODS

An investigation area of about 35 km2 was delimited in Kaliachak Block I (Malda, WB, India), using an As field test (Merck). Located between the Ganges (West) and Bhagirathi river (East) both affected and unaffected areas were examined. In the first field trip March 2002, the main objective was to identify an As anomaly in groundwater and to distinguish it from neighbouring less affected areas. During the second field trip March 2003, the sampling was concentrated on areas with higher As concentrations. Fifty-five household tube and irrigation wells and two dug wells were sampled during the first campaign with an average sampling distance of ⬃600 m. In order to maintain comparable sampling conditions, the second campaign was carried out 12 month later, sampling 71 tube wells with an average sampling distance of 400 m. Before sampling, every well was pumped in order to discharge water standing in the well pipe. In addition to measurement of standard physico-chemical field parameters (pH, conductivity, alkalinity, O2, temperature), the two inorganic As species (AsIII and AsV) were separated in the field, using disposable cartridges filled with selective aluminosilicate adsorbent (Meng et al. 2001). At a pH value of 4–9, As(V) is adsorbed in the cartridge, while As(III) is not. Water samples were filtered (Millex HV 0.45 ␮m) with an aliquot acidified with HNO3 (suprapur) for cation analysis. Major cation and anion concentrations were measured by standard analytical procedures (flame AAS and IC, respectively). 3⫺ 2⫺ During the second year, NH⫹ 4 , PO4 and SO4 were determined using a field photometer on selected samples. Arsenic concentrations and other trace elements (including Fe and Mn) were measured by means of HR-ICP-MS (Axiom, VG Elemental). A small number of samples (n2002 ⫽ 29, n2003 ⫽ 25) were selected for analyzing ␦34SSO . Because 4 of the very low SO42⫺, higher amounts were necessary in order to get the necessary amount of 34 BaSO4 for the determination of ␦ SSO . Therefore, cartridges filled with anion exchange resign 4 were used to enrich SO42⫺ from several liters, following the procedure recommended by the Carmody et al. (1998). In the laboratory, sulphate on the resign was flushed with a KCl solution and precipitated as BaSO4. Sulphur isotope ratios were measured in continuous flow technique using an IRMS coupled on-line with an element analyser (Optima, Micromass). Manual drilling (28 m depth) realized by local people and one rotary drilling (60 m) by Hydro & Geosurvey Consultants Ltd. enabled sampling of aquifer sediments in the affected area. Samples were collected each meter and after drying transported and stored in sealed plastic bags. Mineral composition and trace element concentrations in the bulk material were determined by means of standard XRDA (Kristalloflex D500, Siemens) and ED-XRFA (Spectrace 5000, Atomica). Total carbon, sulphur contents and inorganic carbon were measured using a Carbon-Sulphur-Analyser (CSA 5003, Leybold Heraeus), and a Carbon-Water-Analyser (CWA 5003, Leybold Heraeus). Organic carbon content was evaluated by difference between total and carbonate carbon. In order to investigate the mineral speciation of As in the aquifer sediments, ten samples from different depth intervals of both drillings were investigated by means of ␮-synchrotron-XRFA. For this, 4

double side polished sections of about 100 ␮m thickness were prepared from the dispersed sample material embedded in Epotek resin. A main advantage of the ␮-synchrotron XRFA as compared to current site resolving, non destructive analytical methods is its low detection limit, allowing the quantification of element concentrations in the low mg/kg range. However, it should be mentioned that due to matrix effects, a the different response depth for each element and a varying section thickness on a micron scale, only semi-quantitative results could be obtained. The analyses were carried out at the Angstrom-Source of the Forschungszentrum Karlsruhe, Germany (ANKA).

3

RESULTS

3.1 Stratigraphy and geochemistry of the aquifer sediments As part of the Bengal Delta Plain, the aquifer in the working area is build up by sequences of alluvial fining upward strata, deposited by the precursor of Ganges-Brahamaputra river during Late Pleistocene and Holocene time (Umitsu 1990). Local data about the architecture of the aquifer in the AOI (Kaliachak Block I) were not available. However, a previous groundwater survey in the Harishandrapur Block I (North of the AOI) in Malda suggest a Holocene aquifer with sands of varying grain size with intercalated clay layers and the presence of a deeper aquifer between 150–200 m b.g.l. (IIT Kharagpur, unpub.) Two drillings (28 m, 60 m) shed light on stratigraphy and material of the aquifer (Fig. 1). A continuous sandy aquifer is extending at least up to 60 m depth with changing grain size and thin intercalated silt and clay layers. As typical crossbeded fining upward strata, the clay beds are considered to be cut in relatively short distance by a new fining-upward sequence. Therefore, those fine grained sediments may act locally as aquitard, but play possibly not a major role on regional scale. The aquifer sediments are of greyish colour and consist mainly of quartz, feldspar and mica. Organic carbon and sulphur content throughout the core are relatively low (0.06%, 0.01%), but slightly enriched in the clay (up to 0.3%, 0.02%). The As content in the aquifer sediments is diffusely distributed and with 3–12 mg/kg is within the characteristic range of alluvial sediments. Intercalated silty clay layer may be slightly enriched in As (up to 12 mg/kg). Sediments from the 28 m drilling show a good correlation of As with Fe and Mn in sediments (As-Fe: r ⫽ 0.88; AsMn: r ⫽ 0.79), while in samples from the 60 m drilling such a covariance was not found (As-Fe: r ⫽ 0.37; As-Mn: r ⫽ 0.36). The distribution of selected main and trace elements was investigated on a microscopic scale by means of ␮-synchrotron XRFA, by carrying out line and area scans and recording the concentrations of K2O, CaO, TiO2, MnO, Fe2O3, V, Cr, Ni, Cu, Zn, and As in several hundred to over thousand individual points in each of the thin sections. Results obtained on different samples are consistent and indicate that As concentrations are generally low, with up to several tens of mg/kg. Considering the huge volume of the collected data, and because mostly no univocal correlation was found between the As content and the concentration of other elements, a multivariate statistical approach was applied to throw light on the association of As with different minerals. Individual analytical data points were grouped according to their compositional similarities into chemically homogenous entities, using the k-mean clustering technique (Statistica’99, StatSoft, Inc., USA). Before data processing, each variable was standardized to give an average equal to 0 and standard deviation of 1. As a typical example, the results of the clustering for sample D23 (coarse sand, 6 ppm As bulk; Fig. 1) are showing the mean element concentrations for each of the groups presented in Figure 2a. The statistical analysis defined six groups with distinct geochemical characteristics, each of them representing different mineral species. Based on the relative abundance of the elements, in agreement with the results of optical microscopy and XRD analysis, the clusters can be associated with the following minerals:

• •

Cluster I: background (Epotek resin); Cluster II: muscovite/sericite (high contents in K2O and low in all other elements); 5

Figure 1. Litholog of the 28 m manual drilling (left) and 60 m rotary drilling, including As, Fe and Mn concentrations in solid material along the depth profile. In the lithological column, predominant grain size is visualized by extend of the beds, reaching from silty clay (C), fine sand (fS) up to coarse sand (cS).

Figure 2. (a) Mean standardized element concentrations for each cluster, calculated with the k-means method. (b) Distribution of the As concentrations in a selected thin section of alluvial sediment in relation to the spatial distribution of the clusters. Clusters are in 100 ␮m distance.

6

• • • •

Cluster III: calcite and/or plagioclase feldspars (high contents in CaO and low in all the other elements analysed); Cluster IV: show a similar element pattern as Cluster V, but with generally lower concentration levels. This cluster may represent acquisition points on very thin scales of chlorite in which the beam penetrates in part also into the resin below; Cluster V: chlorite (moderately high contents of Fe2O3, K2O and of trace elements characteristic for mafic minerals like Ti, Ni, Cu, Zn, As, V); Cluster VI: biotite (high contents of Fe2O3, K2O, Ti, Ni, Zn, V, Cr).

The distribution of the As concentrations in relation to the areal distribution of the clusters in the mapped area is shown in Figure 2b. The highest As contents (⬍30 mg/kg) are clearly associated with chlorite, followed by biotite and with only a few ppm of As in sericite, muscovite, calcite and feldspars. It must be considered that the spatial resolution of ␮-synchrotron XRFA may be not sufficient in order to detect very fine grained and disperse distributed secondary Fe and Mn minerals in these samples. Therefore, further investigations are in progress to identify sequences enriched in secondary Fe and Mn minerals. First results of sequential extraction experiments following Keon et al. (2001) confirm the presence of secondary Fe- and Mn phases (data unpubl.). 3.2 Hydrochemistry Based on major anion and cation concentrations the groundwater can be classified as a Ca-MgHCO3 water. Electrical conductivity and the calculated value for total dissolved solids (TDScal) in tube well samples ranged from 400 to 1600 ␮S/cm and from ⬍0.4 to 1.2 g/L, respectively. The values are higher in groundwater taken from open shallow dug wells (up to 3000 ␮S/cm) and up to 2.4 g/L (TDScal), due to evaporation and possible anthropogenic input. Arsenic concentrations in the dug well water are below 10 ␮g/L. 3.2.1 Distribution of As in groundwater With As concentrations of more than 800 ␮g/L, the groundwater in the area can be ranked as highly enriched in As. The depth distribution of the As-concentrations in groundwater (Fig. 3a) is

Figure 3. (a) Depth distribution of Astot in groundwater plotted against the mean filter depth in wells sampled in 2002 (open circle) and in 2003 (crossed circle). (b) Comparison of Astot concentrations in samples taken in 2002 and 2003 in the same tube wells. Dashed lines represent a deviation of ⫾5% from the diagonal.

7

similar to that reported in previous works for the BDP (e.g., BGS & DPHE 2001, Van Geen et al. 2003). Earlier piezometer studies (e.g., McArthur et al. 2004) pointed out that arsenic concentrations in groundwater may change significantly within tens of centimeter depth. Due to the typical local tube well construction with filter length of about 12 feet (⬃3.7 m), groundwater samples taken from tube wells representing the integrated hydrochemical composition along this depth. Deeper pumping stations in Kaliachak were found to be equipped with a 36 feet (⬃11 m) filter tube. Arsenic concentrations ⬎50 ␮g/L are observed generally at depth below 10 m b.s.l., while the highest As concentrations occur in a depth range between 12 and 35 m b.s.l. Even when the number of samples taken from greater depth (up to 85 m b.s.l.) are statistically not representative, no high As concentrations occur at this depth anymore. Nevertheless, As concentrations remain above the Indian drinking water standard (50 ␮g/L) in most of the deeper groundwater samples. While sampling locations varied as a consequence of the different sampling strategies and due to the accessibility of the wells during the two field trips, 19 tube wells were sampled in both sampling campaigns. Details about field data and chemical composition of selected samples are summarized in Table 1. Comparison of these data suggests a stable As value within ⫾5% in 14 of the 19 sampled tube wells. In the remaining 5 wells As contents deviate significantly, being consequently higher during the second sampling campaign (Fig. 3b). In one of them, groundwater was found to be below 50 ␮g/L during 2002 (22 ␮g/L As), while in 2003 the As content (151 ␮g/L) strongly increased. This indicates a local increase of the As concentrations within twelve months. Statistical evaluations of large data sets in previous studies (Fazal et al. 2001, Van Geen et al. 2003) led to the conclusion that a slightly positive correlation exists between As content and the age of the wells. However, considering only samples from wells tapping the depth range where high As concentrations in groundwater occur, no significant correlation between the year of well construction and As content has been found (data not plotted). An IKONOS satellite image of the investigation area is shown in Figure 4a. Villages are outlined from surrounding agricultural areas (mainly paddy fields, wheat fields and mango plantations). Geomorphology reveals the presence of oxbow channels. A N-S heading channel (which is hard to follow northward due to extended settlements) represents most likely an abandoned channel of the recent Bhagirathi river. Two parallel W-E heading oxbow channels are related to an eastward flowing tributary, which recently meets the Bhagirathi river at about 3 km south of the map shown in Figure 4. The spatial distribution of the As contents observed in groundwater samples taken in 2002 (open circles) and 2003 (crossed circles) are shown in Figure 4c. Clearly, the As level exceeds the current Indian drinking water standard of 50 ␮g/L in large parts of the area. Large unaffected areas are only found along the national highway N.H.34 and to the north. To the south and east, the groundwater contains high As concentrations (⬎50 ␮g/L), including several small areas where tube wells with high As contents are concentrated. As mentioned above, arsenic concentrations in tube well water represent the integrated value along the filter depth of 12 feet. Deeper insight into the spatial relationship of arsenic in groundwater results by the interpolation of the As concentrations found during the more detailed sampling in 2003 (Fig. 4b). Though drillings suggest a homogenous and undisturbed aquifer up to at least 60 m depth (Fig. 1), only samples taken from the affected depth range between 12–35 m are considered. Keeping in mind the assumption on a homogenous aquifer in this depth range, a contour plot renders only a simplified image of the real situation. For instance, comparing Figure 4c with 4b, the closed “unaffected” area in the NW (crossed pattern) may include in reality some spots with increased As concentrations, nevertheless this area may still be considered as relatively unaffected. The purpose of Figure 4b is to show the distribution of the main “hot spots” and to reveal a possible pattern in their apparent patchy and irregular distribution. Indeed, the locations of the highest affected wells are noticeably matching the course of old river channels. In Mosimpur the area with high arsenic groundwater is overlying or is close to the old river channel of the Bhagirathi river and seems to extend southward following the course of this oxbow. Furthermore, the water taken from two tube wells at the northern border of the E-W oxbow channel has the second highest As concentration (N of Sayedpur-Makulpur, and S of Jalalpur). Because this oxbow is hardly noticeable further to the West, it is difficult to find any relationship with the high As wells in the SW of the area. 8

9

2754672 2754672 2752107 2752107 2752776 2752776 2754466 2754466 2755653 2755653 2754904 2754904 2754531 2754531 2753898 2753898 2754459 2754459 2752798 2752798 2751870 2751870 2751554 2751554 2752114 2752114 2755125 2755125

610779 610779 605130 605130 605829 605829 609099 609099 609833 609833 610850 610850 611271 611271 611185 611185 609458 609458 608145 608145 606372 606372 607322 607322 608742 608742 610296 610296

29 29 20 20 18 18 23 23 21 21 20 20 26 26 30 30 20 20 18 18 27 27 25 25 23 23 20 20

24.9 24.2 – – 27.0 – 28.2 27.0 27.0 – 28.6 25.3 26.8 27.0 27.0 27.0 27.2 28.0 27.0 25.9 25.7 26.6 26.2 – 27.3 26.0 27.3 26.3

7.2 7.1 7.2 7.2 7.3 7.4 7.2 7.1 7.2 7.1 7.3 7.0 7.2 7.1 7.2 7.1 7.3 7.3 7.2 7.0 – 7.2 7.2 7.2 7.3 7.1 – 7.1

832 833 1480 1135 773 627 1105 885 1160 1070 1652 1601 1105 928 1220 928 796 648 1528 1419 945 944 765 650 840 835 995 878

137 444 5 10 22 151 31 34 36 9 253 367 568 804 569 558 25 23 498 459 462 467 132 118 169 139 205 256

269 446 4 10 17 129 24 33 34 8 225 301 507 746 540 452 23 23 402 384 200 342 49 94 64 113 190 235

175 26 1 0 5 22 9 1 2 1 28 66 61 58 29 106 2 0 96 75 262 126 83 24 104 26 15 21

103 117 134 154 59 81 140 117 139 140 118 117 128 120 120 123 96 98 168 178 114 125 88 94 120 121 130 117

29 30 30 43 11 13 33 29 28 30 31 31 37 43 34 43 23 24 45 43 32 41 22 24 28 26 33 27

16 22 126 40 33 31 36 25 45 46 14 20 22 17 23 27 5 7 46 54 26 28 15 15 17 17 36 23

4.1 5.2 5.5 5.5 3.7 4.7 6.4 5.7 4.9 5.5 2.7 6.0 5.4 6.4 5.4 6.0 4.1 5.3 10.9 14.1 4.8 5.1 3.6 3.6 4.7 5.0 11.6 9.8

1.1 ⬎2.2 0.0 0.5 0.6 1.0 0.9 0.2 0.0 1.2 0.2 ⬎2.2 ⬎2.2 1.7 ⬎2.2 – 0.0 0.3 2.1 ⬎2.2 ⬎2.2 ⬎2.2 1.0 1.5 2.1 – 1.9 1.5

4.25 6.92 0.23 1.41 0.33 0.22 0.65 0.69 0.45 1.22 8.10 13.1 7.70 9.79 8.77 11.9 0.28 0.62 0.87 2.93 0.84 7.23 0.95 7.10 0.75 6.25 0.89 5.78

0.41 0.47 0.59 1.10 0.52 0.60 0.76 0.68 0.93 0.95 0.67 0.52 0.42 0.19 0.45 0.48 0.49 0.53 1.71 1.97 4.71 0.88 2.54 0.85 4.06 0.76 3.58 0.99

0.55 0.58 0.40 0.58 0.21 0.25 0.46 0.37 0.58 0.59 0.51 0.49 0.66 0.84 0.56 0.59 0.33 0.35 0.72 0.71 0.67 0.66 0.35 0.34 0.45 0.42 0.64 0.59

0.24 0.25 0.17 0.23 0.11 0.12 0.24 0.20 0.25 0.23 0.29 0.32 0.39 0.38 0.33 0.32 0.16 0.17 0.37 0.37 0.29 0.30 0.17 0.21 0.34 0.32 0.22 0.22

503 567 739 647 296 378 484 482 547 561 582 610 670 690 684 634 472 451 694 738 554 543 462 439 577 567 621 561

Eh Asto AsIII AsV Ca Mg Na K NH4 Fe Mn Sr Ba HCO3 pH ␮S/cm ␮g/L ␮g/L ␮g/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L

W012002 W012003 W022002 W022003 W032002 W032003 W052002 W052003 W082002 W082003 W092002 W092003 W102002 W102003 W122002 W122003 W132002 W132003 W142002 W142003 W332002 W332003 W342002 W342003 W372002 W372003 W552002 W552003

T. °C

Lat. Long. (UTM, WGS84)

Well

Depth m bgl.

Chemical composition of groundwaters from selected wells, sampled both in Marth 2002 and March 2003.

Table 1.

– – 16.5 14.0 28.5 – – 3.4 17.4 – – 3.0 – 3.2 – – 12.6 – 18.9 – – 12.5 – – – – – 8.21

– 0.73 – 0.10 – 0.09 – 0.07 – 0.08 – 0.76 – 0.99 – – – 0.09 – 0.51 – 0.68 – 0.89 – – – 0.24

⬍3 5.0 12.1 24.0 3.8 ⬍3 57.1 38.0 31.0 29.8 ⬍3 ⬍3 ⬍3 14.0 ⬍3 19.0 3.6 3.9 42.4 71.0 – 8.0 ⬍3 ⬍3 ⬍3 9.0 7.2 11.0 0.4 1.0 44.8 60.3 24.6 23.1 85.0 47.7 87.5 69.8 13.2 11.6 49.4 4.5 1.6 1.9 1.2 2.1 93 101 40.2 48.4 0.4 0.5 4.0 4.5 37.3 22.8

␦34S ‰

Cl SO4 PO4 mg/L mg/L mg/L

2.0 2.1 3.9 1.9 ⫺2.0 ⫺0.7 ⫺0.1 ⫺3.4 ⫺3.3 ⫺0.2 ⫺3.3 ⫺1.5 ⫺7.0 ⫺4.1 ⫺4.0 2.5 ⫺5.6 ⫺1.4 ⫺1.2 ⫺3.2 ⫺1.2 4.6 ⫺1.2 3.6 ⫺1.0 ⫺2.0 ⫺0.1 ⫺2.3

Ion. bal. %

Figure 4. Spatial distribution of Astot and NH⫹ 4 concentrations in groundwater of the working area (Kaliachak Block I, Malda District, West Bengal). Coordinates are given in meter (UTM, WGS 84). (a) satellite image (IKONOS Space Imaging Inc.) with location of drilling sites. Surface water appears in black, villages are outlined with white dotted lines (1: Jadupur, 2: Jalalpur, 3: Mosimpur, 4: Sayedpur-Makulpur). (b) Contour plot of As concentrations (interpolation with ordinary linear kriging). Only wells filtering in the main affected depth range (12–35 m) are included. (c) Location of sampled tube wells in 2002 (n ⫽ 55, open circle) and in 2003 (n ⫽ 71, crossed circle). Symbol size is proportional to the observed As concentration. (d) Distribution of the NH⫹ 4 concentrations in 2003 (n ⫽ 54).

3.2.2 Redox-sensitive solutes The As in groundwater is generally dominated by As(III) species, averaging 76% (2002) and 86% (2003) of the total As content. Ninety percent of all samples, including those with the highest Astot concentrations, have shown an As(III)/Astot ratio between 0.6 and 1.0. Organic As species (e.g., MMAA, DMAA) are considered to be negligible based on previous analyses (BGS & DPHE 2000). The high proportion of As(III) is in agreement with a release of arsenic under reducing conditions. The increasing share of As(III) relative to As(V) may suggest a possible trend towards slightly more reducing conditions from March 2002 to March 2003. This scenario is supported by hydrochemical particularities, mainly in respect of the Fe and Mn concentrations, in samples taken in 2002 and in 2003 (Fig. 5). In 2002 high As contents are generally coupled either with high Mn (up to 6.1 mg/L, median: 1.0) or with high Fe concentrations (up to 8.8 mg/L, median: 0.7) (Fig. 5a). The available geochemical data on the sediments (see section 3.1) does not allow to explain the occurrence of groundwater samples with high Mn and low Fe concentrations by high availability of Mn in the 10

Figure 5. Variation of Fetot and Mntot concentrations in groundwater samples collected in 2002 (a) and 2003 (b) Symbol size is proportional to the As contents.

aquifer sediments. Moreover, Mn contents in sediment are relatively low (up to 0.8 g/kg) and are correlating well with the Fe contents (rFe,Mn ⫽ 0.89). We consider that desorption reactions or cation exchange are not sufficient to explain up to 6 mg/L Mn and up to 14 mg/L Fe in groundwater of an aquifer with the observed composition. The missing covariance between the Fe and Mn concentrations in the groundwater more likely can be explained by a specific hydrochemical environment. High Mn and low Fe contents in groundwater may be due to local redox conditions in which only the reduction of Mn but not that of Fe occurs. Due to the stability of Fe-oxihydroxides under Mn(IV)-reducing conditions, we suggest Mn-oxide phases (e.g., birnessite, pyrolite, cryptomelane) as possible source for As (see discussion). While arsenate adsorption in synthetic birnessite is reported to be negligible (Oscarson et al. 1983), 0.1 g of birnessite with an arsenite sorption capacity of 164 ␮M/g (Oscarson et al. 1983, Stollenwerk 2003) may adsorb up to 1230 ␮g As(III), which can be readily mobilized by competitive ion exchange or by the reductive dissolution of birnessite. Based on this we may assume distinct redox zones within the working area under which (though separated in space and/or time) the selective reduction of Mn(IV) and Fe(III) occur. Samples enriched in both As and Fe were found in 2002 in the village Mosimpur, suggesting an environment with lower redox potential in this area. In samples taken in 2003 the Fe concentrations increased significantly (up to 13.1 mg/L, median: 3.1), while Mn rarely exceeded 1 mg/L (max. 2.0 mg/L, median 0.7). This suggests a general decrease in the redox potential between the two sampling campaigns, reaching Fe-reducing conditions coupled with a slightly increase in the As contents. Similar to 2002, the highest Fe concentrations occurred in Mosimpur. The lack of high Mn concentrations in most of the Fe-rich groundwater samples may be due to flushing of high Mn groundwater or to the precipitation of non-oxidic Mn-phases. Calculation of saturation index for rhodochrosite (MnCO3) using the hydrochemical code PHREEQC results in positive values for sampling in 2002 (SIrhod. ⫽ 0.05⫺1.44, median: 0.77; estimated Eh ⫽ 0 mV) and therefore confirms the tendency of manganese to precipitate as rhodochrosite. The amount of HCO⫺ 3 in the groundwater calculated from alkalinity under near neutral pH conditions varies between 300 and 800 mg/L. Additional to the dissolution of carbonates, HCO⫺ 3 forms during the degradation of organic matter. Under anoxic conditions the production of HCO⫺ 3 ⫺ is related to the presence of different electron acceptors (NO⫺ 3 , Mn(IV), Fe(III), SO4 ) increasing in groundwater the concentrations of species like NH4, Mn(II), Fe (II) or H2S. In the study site ⬃70% of all samples have a HCO⫺ 3 /(Ca ⫹ Mg) [meq/L] ratio of ⬎1 (up to 1.4), which would suggest that at least a part of HCO⫺ 3 originates from reduction of organic matter. 11

The presence of dissolved ammonium at concentrations greater than 1 mg/L in groundwater either as direct input or due to ammonification implies an organic N source. High NH⫹ 4 contents in groundwater are more frequently encountered in villages than in agricultural areas (Fig. 4d) and NH⫹ 4 correlate well with As concentrations (r ⫽ 0.73). Consequently, fertilizer may be considered to play a minor role as a source for N in the aquifer, possibly due to its rapid uptake by agricultural crops. Other conceivable sources are natural organic matter distributed disperse in the sediments or enriched in some lithologies, from where it may be mobilized during the microbial degradation of organic matter (McArthur et al. 2004). It may be also introduced into the aquifer from nutrient rich sewage which locally may percolate from pits. Such kind of an input may be enhanced by both, intense groundwater exploitation and along unsealed wells. The spatial clustering of several wells with high ammonium contents like in the area of Mosimpur and along the E-W stretching oxbow channel (Fig. 4d) may indicate a relative widespread and therefore natural source for the organically derived N. In contrary, in spot like occurring high As sites, like in the south of Jadupur and center of Jalalpur a more local input may be assumed (Fig. 4b, 4c). The unusually high concentrations of Cl⫺ (48 mg/L, 72 mg/L; median2003: 5.6 mg/L) support the assumption for a local anthropogenic input in these wells. Here, the input of nutrient rich sewage water may be the main driver for accelerated microbial activity. The sulphate content in groundwater ranges from ⬍1 mg/L and 90 mg/L (median: 5 mg/L). At least in part, anthropogenic contamination may be considered as a possible cause for high sulphate concentrations in some of the wells. In order to use the ␦34SSO⫺ values as an indicator for sulphate 4 reduction, sulphur isotope ratio measurements were conducted on a limited number of samples (n2002 ⫽ 29, n2003 ⫽ 25). Due to the preferential consumption of the 32S isotopes by sulphate reducer bacteria, under closed system conditions (limited sulphate pool) while reduced sulphur species (H2S) are depleted in 34S, residual sulphate will be progressively enriched in the 34S. Initiated under strongly reducing conditions (Eh ⬃ ⬍⫺150 mV; Postgate 1984), the fractionation of the sulphur isotopes depends on the rate of sulphate reduction, which further depends on the quality of organic matter and to a lesser extent on the concentration of the available sulphate (Crossman & Desrocher 2001.). According to Figure 6, no direct relationship can be found between the ␦34SSO values and the 4 As content of groundwater. High As concentrations (⬎50 ␮g/L) occur with sulphate both depleted 34 and enriched in S. However, high As concentrations (As⬎200 ␮g/L) seem to be coupled with highly varying ␦34SSO values (1.5–25‰) and with relatively low SO42⫺ contents (⬍14 mg/L). 4

Figure 6. ␦34SO4 values as a function of sulfate concentration (left: linear scale, right: logarithmic scale), in samples taken in 2002 (n ⫽ 29) and in 2003 (n ⫽ 25). Samples are classified according to their As concentrations into low (0–50 ␮g/L), intermediate (50–200 ␮g/L) and high (⬎200 ␮g/L). Grey shaded area indicates the range of sulphur isotope composition in monsoon rain and river water (Jacks et al. 1994, in Zheng et al. 2004).

12

Several samples have ␦34SSO values above the isotope composition of sulphur in the monsoon rain 4 (10‰) and in the water of the Ganges-Brahamaputra river system (⬃3–12‰; Jacks et al. 1994, in Zheng et al. 2004; Fig. 6). The hinted general trend in Figure 6 suggests sulphate reduction as an ongoing process in groundwater at some of the sampling sites. Arsenic is possibly not precipitated along with iron-sulphides, which may be due to the deficiency of dissolved iron or of sulphate (Zheng et al., 2004). High As water coupled with low ␦34SSO values may be explained by a reoxidation 4 of sulphides to sulphate (Zheng et al. 2004). However, the lack of concrete data on dissolved oxygen and Eh does not allow the verification of such a scenario. Unclear is also the reason for the occurrence of the three samples with extremely high ␦34SSO values (⬎50). 4

4

DISCUSSION AND CONCLUSIONS

The patchy distribution of the high As concentrations in groundwater seems to be related to the presence of a small scale redox zonation in the aquifer. Taking high amounts of dissolved Mn and Fe as well as high ␦34Ssulphate values as indirect indicators for low redox conditions, the release of As can be discussed in terms of progressively decreasing redox potentials at which the reduction of Mn(IV), Fe(III) and SO42⫺ takes place. The mobilization of As at relatively high redox levels would imply the presence of As-bearing Mn-oxides in the aquifer (e.g., birnessite, cryptomelan). This should be taken into consideration when the role of Mn-oxides in the sorbtion/desorption of As is discussed (Stüben et al. 2003, Akai et al. 2004, McArthur et al. 2004). Evidences for stronger reducing conditions as such necessary for sulphate reduction were also found, as indicated e.g., by the local occurrence of low sulphate contents and/or high ␦34Ssulphate values coupled with high As concentrations. Arsenic may locally be enriched in SO4⫺reducing groundwater (Fig. 6), possibly due to a lack of sufficient Fe2⫹ or sulphate to be consumed entirely by co-precipitation with Fe-sulphides. Because of the massive mobilization of As from Fe-oxihydroxides under such conditions, the slight decrease of As concentration in water due to co-precipitation with Fe-sulphides may be completely masked. Though the presence neither of Mn- nor of Fe-oxidic phases could be directly documented by the mineralogical methods used, the presented data and preliminary results of currently running sequential extraction experiments strongly suggests the local presence of these mineral phases. Due to the strong influence of microbial catalyzed oxidation of organic matter on the redox potential (as generally accepted, see e.g., Langmuir 1997), the rapid and small scale change in the hydrochemical environment is mainly caused by interaction with factors which accelerate the degradation of organic matter. Because the redox environment is controlled by the distribution of the amount, availability and quality of natural organic matter within the aquifer they are also the main controlling factors for the mobilization of As. These may be modified by the local input of sewage which percolate from pits or along of macro-channels deep into the subsurface. A closer look at the distribution of highly affected wells suggests a spatial relationship between high As concentrations in groundwater and the presence of oxbow channels. The reason may be found in the lithological composition of the sediments. After the formation of a new river channel, the old channel turn into an oxbow lake in which a higher amount of fine grained sediments and organic matter accumulate. This may favor the degradation of the organic matter accelerating the development of a reducing environment. The occurrence of dissolved ammonium in the groundwater along such channels supports this scenario, but does not exclude also a possible local input of anthropogenic sewage which would favor the development of a reducing environment in parts of the aquifer low in natural organic matter. Certainly, a spatial coincidence of these circumstances would enhance the development of conditions necessary for the release of arsenic into groundwater. Such a scenario may be envisaged for Mosimpur, a densely populated village overlaying the oxbow channel of the Bhagirathi river, where the highest As concentrations in groundwater has been found. Further studies would be necessary to distinguish between sewage related nutrients and naturally occurring organic matter distributed within the aquifer and to estimate the impact of oxbow channels on local groundwater hydrochemistry. 13

Despite of the heterogeneous spatial distribution of redox zones in the aquifer, a slight general trend toward the development of stronger reducing conditions was observed during the time span of 12 month. In 2002, the groundwater hydrochemistry was dominated by Mn-reducing conditions, though also zones with Fe⫺ and SO4⫺ reducing were present. A year later, the groundwater hydrochemistry indicated a shift toward lower redox potentials as it is suggested by a significant increase of the arsenic contents in part of the tube wells sampled during both campaigns. Though the influence of climatic factors such as a difference in annual precipitation cannot be excluded, this difference indicates that the release of As into the groundwater is an ongoing dynamic process. Therefore, even apparently low As wells should be monitored periodically in order to assure a safe supply with non-affected water. Though the deepest wells (40–80 m b.g.l.) sampled during both field campaigns are relatively low in As, they seem to be not suitable for extracting “As-free” water. In the working area, these deep wells are furnished with pumping stations, providing irrigation water for the surrounding agricultural areas. Because of a possible hydraulic connection to shallower stronger affected parts of the aquifer, the intensive groundwater withdrawal would probably accelerate the infiltration of high As groundwater and would locally contaminate deeper parts of the aquifer. Therefore, the installation of deeper tube wells in order to supply the population with safe drinking water should be decided carefully, taking always into consideration the local hydrogeological situation. The aquifer sediments generally show low As-concentrations. Analyses by means of ␮-synchrotron XRF confirmed the diffuse distribution of As on ␮m-scale. Arsenic contents of up to 100 mg/kg were measured only in a few mineral grains. In the samples studied so far, Fe-silicates such as biotite and chlorite were found to be the main As carriers of the aquifer sediments. However, at this stage of the investigations the possibility can not be excluded that extremely fine grained and disperse distributed oxihydroxides are masked in the ␮-XRF analytics by the quantitatively more important and larger Fe-silicates. Even if in low quantity, because of the large reactive surface of the disperse Fe-/Mn-oxihydroxides and their high affinity to adsorb large amounts of As (e.g., Oscarson et al. 1983, Stollenwerk 2003) they must be considered to play a major role in the occurrence of As-rich groundwater hot-spots as recently emphasized also by McArthur et al. (2004). Further investigations are in progress to identify sequences enriched in secondary Fe and Mn minerals.

ACKNOWLEDGEMENTS This work was made possible by a grant offered to one of the authors (FW) by the Deutsche Forschungsgemeinschaft in the frame of a PhD program on natural disasters. We thank also Prof. D. Chandrasekharam (IIT Mumbai) and Dr. D. Chatterjee (Univ. Kalyani) for their support in organisation of the field trips and to Mrs. P. Aggarwal and Mr. B. Nath for their assistance and valuable help during these campaigns. Dr. U. Kramar was of great help in carrying out the analytics with the ␮-synchrotron radiation source.

REFERENCES Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A.H., Imam, M.B., Khan, A.A. & Sracek, O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh: an overview. Applied Geochemistry 19: 181–200. Akai, J., Izumi, K., Fukuhara, H., Masuda, H., Nakano, S., Yoshimura, T., Ohfuji, H., Anawar, H. & Akai, K. 2004. Mineralogical and geomicrobiological investigations on groundwater arsenic enrichment in Bangladesh. Applied Geochemistry 19: 215–230. BGS & DPHE 2001. Arsenic contamination of groundwater in Bangladesh. In D.G. Kinniburgh & P.L. Smedley (eds.), British Geological Survey, Technical Report WC/00/19, London. BES 2002. District Statistical Handbook 2001 – Malda. Bureau of Applied Economics & Statistics, Government of West Bengal, Kolkata.

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Carmody, R.W., Plummer, L.N., Busenberg, E. & Coplen, T.B. 1998. Methods for collection of dissolved sulfate and sulfide and analysis of their sulfur isotopic composition. U.S. Geological Survey OpenFile Report 97–234: 91 pp. Crossman, E.L. & Desrocher, S. 2001. Microbial Sulfur Cycling in Terrestrial Subsurface. In J.K. Fredrikson & Fletcher: M.F. Subsurface Microbiology and Biogeochemistry: 219–248. New York: Widely-Liss. IIT Kharagpur (unpublished). Semi-detailed ground water survey in Malda & W. Dinajpur Dts, Harishandrapur (I) Block. Rural Development Centre, Indian Institute of Technology Kharagpur. Jacks, G., Sharma, V.P., Trossander, P. & Aberg, G. 1994. Origin of sulphur in soil and water in a Precambrian terrain. India. Geochem. J. 28: 351–358. Keon, N.E., Swartz, C.H., Brabander, D.J., Harvey, C. & Hemond, H.F. 2001. Validation of an arsenic extraction method for evaluating mobility in sediments. Environ. Sci. Technol. 35: 2778–2784. Langmuir D. 1997: Aquous environmental geochemistry. 600 pp. New Jersey: Prentice Hall. McArthur, J.M., Ravenscroft, P., Safiullah, S. & Thirlwall, S.M. 2001. Arsenic in groundwater: testing pollution mechanisms for sedimentary aquifers in Bangladesh. Water Resources Research 37(1): 109–117. McArthur, J.M., Banerjee, D.M., Hudson-Edwards, K.A., Mishra, R., Purohit, R., Ravenscroft, P., Howarth, R.J., Chatterjee, A., Talukder, T., Lowry, D., Houghton, S. & Chadha, D.K. 2004. Natural organic matter in sedimentary basins and its relation to arsenic in anoxic ground water: the example of West Bengal and its worldwide implications. Applied Geochemistry 19: 1255–1293. Meng, X., Korfiatis, G.P., Christodoulatos, C. & Bang, S. 2001. Treatment of arsenic in Bangladesh well water using a household co-precipitation and filtration system. Water Research 35(12): 2805–2810. Nickson, R.T., McArthur, J.M., Burgess, W.G., Ravenscroft, P., Ahmed, K.M. & Rahman, M. 1998. Arsenic Poisoning of Bangladesh Groundwater. Nature 395: 338. Norra, S., Aggarwala, P., Berner, Z., Wagner, F., Stüben, D. & Chandrasekharam, D. (submitted). Impact of irrigation with As rich groundwater on soil and crops: a geochemical case study in Maldah District, West Bengal. Oscarson, D.W., Huang, P.M. & Liaw, W.K. 1983. Kinetics of oxidation of arsenite by various manganese dioxides. Soil Science Society of America Journal 47: 644–648. Postgate, J.R. 1984. The Sulphate Reducing Bacteria, 2nd ed. London, Cambridge University Press. Fazal, M.A., Kawachi, T. & Ichion, E. 2001. Validity of the Latest Research Findings on Causes of Groundwater Arsenic Contamination in Bangladesh. International Water Resources Association Water International 26 (2): 380–389. Ravenscroft, P., McArthur, J.M. & Hoque, B. 2001. Geochemical and palaeohydrological controls on pollution of groundwater by arsenic. In: Arsenic Exposure and Health Effects IV., W.R. Chappell, C.O. Abernathy, R. Calderon (eds.): 53–77. Oxford, Elsevier Science Ltd. Smith, A.H., Lingas, E.O. & Rahman, M. 2000. Contamination of drinking-water by arsenic in Bangladesh: a public health emergency. Bulletin World Health Organisation 78(9): 1093–1103. Smedley, P.L. 2003. Arsenic in groundwater – south and east Asia. In A.H. Welch & K.G. Stollenwerck (eds.): Arsenic in Groundwater – Geochemistry and Occurrence , Dordrecht, Kluver. Stollenwerk, K.G. 2003. Geochemical Processes Controlling Transport of Arsenic in Groundwater: A Review of Adsorption. In A.H. Welch & K.G. Stollenwerck (eds.): Arsenic in Groundwater – Geochemistry and Occurrence, 67–100, Dordrecht, Kluver. Stüben, D., Berner, Z., Chandrasekharam, D. & Karmakar, J. 2003. Arsenic pollution in groundwater of West Bengal, India: Geochemical evidences for mobilization of As under reducing conditions. Applied Geochemistry 18(9): 1417–1437. Umitsu, M. 1990. Late Quaternary sedimentary environments and landforms in the Ganges Delta. Sedimentary Geology 83: 177–186. Van Geen, A., Zheng, Y., Versteeg, R., Stute, M., Horneman, A., Dhar, R., Steckler, M., Gelman, A., Small, C., Ahsan, H., Graziano, J.H., Hussain, I. & Ahmed, K.M. 2003. Spatial variability of arsenic in 6000 tube wells in a 25 km2 area of Bangladesh. Water Resources Research 39(5): 1140–1155. Zheng, Y., Stute, M., van Geen, A., Gavrieli, I., Dhar, R., Simpson, H.J., Schlosser, P. & Ahmed, K.M. 2004. Redox control of arsenic mobilization in Bangladesh groundwater. Applied Geochemistry 19: 201–214.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Genesis of arsenic contamination of groundwater in alluvial Gangetic aquifer in India S.K. Acharyya & B.A. Shah Department of Geological Sciences, Jadavpur University, Kolkata, India

ABSTRACT: Arsenic pollution in groundwater mainly affects major parts of the GangaBrahmaputra delta in West Bengal and Bangladesh, as well as, parts of narrow entrenched lowermiddle sections of the Gangetic floodplains. Arsenic adsorbed on hydrated ferric oxides (HFO) was preferentially entrapped in organic rich deltaic Holocene sediments and less frequently in its floodplains. Severe reducing condition that developed later mobilized arsenic to groundwater mainly in the deltaic domain. The sediment cover on the Pleistocene uplands in the Bengal Basin and the interfluve Ganga plain are free of arsenic problem. Arsenic is mobilized to groundwater by bio-mediated reductive dissolution of HFO. Strong reducing nature of groundwater in the Bengal Basin and parts of affected flood plains is shown by high concentration of iron (⭐9–36 mg/L), which is generally low (⬍1 mg/L) in the Ganga alluvial plain upstream of the Bengal Basin indicating that groundwater is not adequately reducing in nature to mobilize arsenic. 1

INTRODUCTION

Arsenic contamination of groundwater affects extensive low-lying deltaic areas in the Bengal Basin, located mainly to the east of Bhagirathi river in West Bengal (India) and Bangladesh, and parts of narrow entrenched middle Ganga floodplain in parts of Bihar, Jharkhand and eastern Uttar Pradesh (UP; Fig. 1). The upper permissible limit of arsenic in potable water is set at 50 ␮g/L in

Figure 1. Map showing Quaternary basins in northern parts of the Indian subcontinent. 1. Ganga Alluvial Plain. 2. Bengal Basin. Abbreviated localities: A – Allahabad, BX – Bauxar, C – Chhapra, B – Balia, P – Patna, BG – Bhagalpur.

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India, whereas, recommended limit is 10 ␮g/L by WHO. Arsenic concentrations in tube-well water exceed 50 ␮g/L limit manifold in parts of the affected area and several cases of arsenical diseases have been recorded, particularly from the pandemically affected parts of the Bengal Basin. Arsenic is also recorded to have entered into food chain from areas irrigated with arsenic laced groundwater (Sanyal & Naser 2002). There is wide variability in level of arsenic concentrations in space and time, with high values restricted to some ‘hot-spots’, whereas, arsenic safe zones are also located within broadly affected areas. Regular monitoring of arsenic level in tube-wells is thus a necessity within arsenic affected and potentially risk areas. Arsenic contamination also affects isolated patchy areas, some of which are located on arsenic enriched acid magmatic rocks. In the affected areas in northern parts of the Proterozoic Dongargarh rift zone, Chhattisgarh, tube-well water is contaminated but dug-wells are generally arsenic free except in Kaurikasa area, where arsenic enriched regolith and soil are exposed. In this area even some dug-wells are arsenic polluted (Acharyya 2002). 2

ARSENIC AFFECTED FLOOD-DELTA PLAINS IN WEST BENGAL

Pandemic arsenic contamination in the Bengal Basin is essentially confined to the low-lying Ganga-Brahmaputra flood-delta plain located downstream of the Rajmahal Hills. The low-lying flood-delta basin of the Ganga, Bhagirathi, Jamuna and old Bhramhaputra rivers is entrenched and incised on the Pleistocene terraces during the low stand setting of the terminal Pleistocene period. The Pleistocene upland (and hilly belt of older rocks) flank the western margin of the Bengal

Figure 2. Landforms and decompositional environments in the Bengal Basin. Legend: 1. Hills of older rocks. 2. Laterite/Ferrisol Pleistocene upland. 3. Older Alluvial Plain. 4–7. Younger Alluvial Plain. 4. Recent Flood and Deltaic Plain. 5. Interdistributary Swamp. 6. Tidal Swamp. 7. Trippera Surface. 8. Location of boreholes along section lines. Abbreviations: Bl-Balagarh, B-Barind, C – Calcutta, Dk-Dhaka, Gh – Ghetugachi, Kh – Khulna, M – Madhupur tract, Md – Malda town, L – Lalmai Hills, R – Rajmahal Hills, TC – Tripura-Chittagong Hills.

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Basin in West Bengal (India), whereas, those at Barind and Madhupur areas mark the northern and central parts of the basin in Bangladesh. The Bengal Basin is flanked to the east by the Tertiary hills (Fig. 2). The Pleistocene sediments on incised up-lands are oxidized to iron-stained, heavy mineral deficient sand and brown-orange stiff clay. These were well flushed by groundwater over longer period and are arsenic free. The Holocene subsurface units beneath the arsenic affected younger delta plain of the Bengal Basin are tentatively subdivided into three units (Acharyya et al. 2000) and the classification is also adopted in Bangladesh (Uddin & Abdullah 2003). The ‘basal unit’ of late Pleistocene – early Holocene sequence comprising gravelly sand above the disconformity was deposited as incised channel fills or fluvio-deltaic sand of proto-Ganga-Bhagirathi-Bharahmaputra rivers around 18,000 to 10,000 yr. B.P., when sea level rose rapidly. The basal gravelly sands are generally micaceous, heavy mineral rich and coarsening northward from fine to medium and coarse sand. The ‘basal unit’ of the Holocene channel fill sands is generally free of arsenic problem. The Ganga-Brahmaputra delta sedimentation of the ‘middle unit’ was induced over large area beginning around 10,000 yr. B.P., when rapid rise of sea level led to back-flooding and overtopping of the low stand entrenched channels and the oxidized late Pleistocene surfaces. Continued high stand, during most of the early-mid Holocene period flooded partly sedimented valley courses and converted their lower and adjacent parts to fluvial marshes, lagoons and estuary. The ‘middle unit’ comprises silt dominated mud and fine sand that commonly contain wood and other plant fragments, shell clasts and marine organic remains. High sediment load from the rapidly eroding Himalayas competed with rapid sea level rise to enforce continued sluggish deltaic sedimentation. The lenticular muddy sand bodies generally form numerous transient distributary channels (Acharyya et al. 2000). Most arsenic contaminated tube-wells mainly tap aquifers in the ‘middle unit’, which were very poorly flushed by groundwater due to their deltaic setting. Thus any arsenic released from these sediments following burial accumulated in groundwater (Fig. 3).

Figure 3. Profiles and correlation of Holocene sediments. A: Section AB (location shown in Fig. 2) across GangaBhagirathi delta in West Bengal. B: Section CD (location shown in Fig. 2) across Jamuna flood plain and Ganga delta in Bangladesh. Modified after Umitsu (1993). 1–3 are broad Holocene stratigraphic units referred in the text.

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The top mud facies of the ‘upper unit’ presently cap sandy sequences throughout the Bengal Basin. These are deposited during rapid sea level rise since 7000 yr. B.P., when sea level reached higher than present level and the southern parts of the Ganga delta was invaded by tidal mangrove and encroached by the Bay of Bengal. There was extensive development of marine and fresh water peat during 7000 to 2000 yr. B.P., within the clayey sequence of the ‘upper unit’, which were mainly confined to southernmost part of the basin. The ‘upper unit’ sediments are also enriched in arsenic but there is minor development of aquifers that are free of saline water within this sequence.

3

ARSENIC AFFECTED AREAS IN MIDDLE GANGA FLOOD PLAIN

Narrow tracts of arsenic affected areas are recorded recently from middle Ganga flood plain from parts of Bihar, Jharkhand and eastern UP, and it is apprehended that arsenic contaminations would affect wide regions of the Ganga alluvial plain (Chakraborti et al. 2003). The affected areas are confined to the Newer Alluvium (Holocene) within narrow entrenched active flood plain. Major parts of the Ganga alluvium interfluve upland plains are unlikely to be affected according to us (Acharyya & Shah 2004a). Sedimentation in these entrenched flood plains was also influenced by sea level fluctuation during the Holocene, causing increased aggradation and forming fluvial swamps (Singh, 2001). The arsenic bearing aquifers are located close to the Ganga and western down faulted side of the Ghaghra river channel-floodplain in parts of Bauxar, Bhojpur and Balia districts, but extensive area exposing or having oxidized Older Alluvium (Pleistocene) at shallow depths to the east of Ghaghra river and along the northern bank of the river Ganga in Chhapra and Baishali districts are unaffected (Fig. 1). Over 80% of the Ganga alluvium plain is represented by older Alluvium interfluve upland (cf. Kumar et al. 1996), which is also unaffected. The Son alluvium plain close to its confluence with the Ganga is also free of arsenic contamination. Our study in this area is under progress.

4

SOURCE AND RELEASE OF ARSENIC IN GROUNDWATER

No specific sources of arsenic could be identified for the Ganga-Brahmaputra river system and potential sources are located both in the Himalayas and Peninsular India. Contrary to claim otherwise, our mineralogical studies indicate that arsenic rich pyrite (Fig. 4) or arsenic minerals are rare or absent in the aquifers from affected areas in West Bengal. However, rare presence of biogenic pyrite is recorded in reducing environment often in association of degraded plant remains (Acharyya 2001, Acharyya & Shah 2004b). These have acted as sinks for and not sources of arsenic. Arsenic contamination is moderate in aquifer sediments from affected and adjacently located arsenic safe areas in West Bengal (Ghetugachhi and Baruipur area; Fig. 2). Studies on drill cores of aquifer sediments from arsenic polluted and safe zones located within overall arsenic affected areas in West Bengal reveal common occurrence of iron-coated quartz and clay (illite) grains, ironmanganese-siderite, magnetite and biotite/chlorite, which are arsenic bearing (Acharyya 2001, Pal et al. 2002). Sludge samples from Balagarh block (Fig. 2), Hoogly district, contain minor fractions of peaty wood fragments (Fig. 5) within which arsenic is locked in authigenic and frambroidal pyrite (Acharyya & Shah 2004b). Aggregates of Fe-Mn-siderite concretions often having biogenic colony like structure and frambroidal pyrite have been found in aquifer sediments from Balagarh and Ghetugachhi area in Hoogly and Nadia districts respectively (Pal et al. 2002, Acharyya & Shah 2004b). Arsenic release by oxidation of pyrite has been disapproved in general, because pyrite is nearly absent in the affected aquifer sediments and sulfate concentrations are very low in affected groundwater. Biomediated reductive dissolution of hydrated ferric oxides (HFO) that occur mainly as coatings on sediment grains and corresponding oxidation of sedimentary organic matter is regarded as the main mechanism, which mobilizes arsenic to groundwater from aquifer sediments (Bhattacharya et al. 1997, Nickson et al. 1998, 2000, Kinniburgh & Smedley 2001). Arsenic sorbed 20

Figure 4. Biogenic pyrite (marked P) in a carbonaceous shale clast. Pyrite growth often follows grain boundary. Arsenic bearing nature of pyrite revealed by SEM-EDX scan, Balagarh-Sripur area, Balagarh.

Figure 5.

Degraded woody fragment where pyrite is replacing cell structures.

in discrete phases of Fe-Mn-oxyhydroxide was preferentially entrapped in argillaceous and organic rich early-mid Holocene deltaic sediments and the Holocene floodplain sediments. Recent studies has established that iron-rich groundwater is produced by the activities of anaerobic heterotropic Fe3+ reducing bacteria (IRB), which preferentially reduce and dissolve least crystalline discrete 21

phases of HFO, with consequent release of its sorbed arsenic and other trace elements to groundwater. Ferrous ion, released by IRB from Fe-bearing mineral phases or HFO sediment coatings possibly reacted with abundantly present bicarbonate in groundwater to precipitate siderite concretions, which grew around sediment grains and/or centers of IRB colonies (Acharyya & Shah 2004b). Reduction of HFO is common and intense in affected aquifer in the Bengal Basin and parts of Ganga floodplain. This is demonstrated by high concentration of dissolved Fe (⭐9–36 mg/L) in arsenic contaminated groundwater (Acharyya et al. 1999, Acharyya & Shah 2004a,b). Under stronger reducing condition and in presence of organic carbon, sulfur reducing bacteria (SRB), instead would precipitate pyrite which would co-precipitate arsenic from groundwater. The chemistry of arsenic affected tube-well water is so far based on study of mixed samples from the entire column. An inflatable packer-stradle-pump assembly was used by us (Guha et al. in prep.), to test chemical characteristics of aquifer water from a specific depth. Interpretation of sediment and water analysis indicates that iron-reducing condition develop at several levels releasing arsenic from sediments to the groundwater. Although arsenic is present in the sediments throughout the entire depth of boreholes, it is not released under nitrate and sulfate reducing conditions. Clayey lenses in the aquifer create low permeability zones preventing electron acceptors like nitrate and sulfate to reach these levels where iron reducing conditions and release of arsenic prevail. The presence of tritium, high 14C (⬃80–112 pMC) and ␦18O values (⫺3.5 to ⫺5.5 ‰) in shallow aquifer groundwater in the Bengal Basin (Shivanna et al. 2000, Agrawal et al. 2000) indicates continuing recharge from local rain, surface and floodwater. Extensive groundwater irrigation has accelerated flow of groundwater that brought dissolved degraded organic matter in contact with HFO bearing sediments, enhancing reduction process and triggering release of arsenic (Acharyya 2001). Arsenic contamination in groundwater in alluvial aquifer is typically confined to organic rich fluviodeltaic sediments e.g. Ganga-Brahamaputra delta in the Bengal Basin (Nickson et al. 1998, 2000, Kinniburgh & Smedley 2001), Red and Mekong River deltas in Vietnam (Berg et al. 2001). Major parts of the Ganga alluvial plain is also subjected to equally intensive groundwater irrigation, but bulk of the interfluve upland in the Ganga plain corresponding to Older Alluvium (Pleistocene) (cf. Kumar et al. 1996) are unaffected by arsenic contamination. However, Holocene alluvium within narrow entrenched floodplain in the parts of lower-middle Ganga plain in Bihar, Jharkhand and UP are arsenic contaminated. Arsenic affected local pockets may also occur in northern fan areas as recorded from terai region in Nepal (Chitrakar & Neku 2001, Bhattacharya et al. 2003). 5

CONCLUSIONS

Pandemic arsenic pollution in groundwater mainly affects low-lying entrenched flood-delta plains of the Bengal Basin covering parts of West Bengal (India) and Bangladesh. The contaminated aquifers are mainly confined to delta sediments deposited around 10,000–7500 yr. B.P., when sea level rose rapidly establishing high stand setting. Arsenic contamination also affects Holocene entrenched floodplain in parts of lower-middle Gangetic plain in Bihar and eastern UP. Sediment cover on Pleistocene uplands in the Bengal Basin as well as, in interfluve uplands in lower-middle parts of the Gangetic floodplain are oxidized and free of arsenic. Pyrite or any other arsenic bearing mineral are nearly absent in aquifer sediments. Arsenic sorbed in phases of iron-oxyhydroxide was preferentially entrapped in organic-rich clayey deltaic sediments and partly in floodplain sediments in lower-middle sections of the Ganga plain. Severe reducing conditions developed later, mainly in the delta sediments and partly within entrenched floodplains, releasing arsenic to groundwater by reductive dissolution of iron-oxyhydroxides. Arsenic contaminated groundwater from the affected areas is thus generally enriched in iron (⭐9–36 mg/L). REFERENCES Acharyya, S.K. 2001. Arsenic pollution in groundwater from lower Ganga plains, Bengal Basin. Indian Journal of Geology 73: 1–19.

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Acharyya, S.K. 2002. Arsenic contamination in groundwater affecting major parts of southern West Bengal and part of western Chhattisgarh: Source and mobilization process. Current Science 82: 740–744. Acharyya, S.K., Chakraborty, P., Lahiri, S., Raymahashay, B.C., Guha, S. & Bhowmik, A. 1999. Arsenic poisoning in the Ganges delta. Nature 401: 545. Acharyya, S.K., Lahiri, S., Raymahashay, B.C. & Bhowmik, A. 2000. Arsenic toxicity in groundwater in parts of Bengal Basin in India and Bangladesh: Role of Quaternary stratigraphy and Holocene sea level fluctuation. Environmental Geology 39(10): 1127–1137. Acharyya, S.K. & Shah, B.A. 2004a. Risk of arsenic contamination in groundwater affecting the Ganga alluvial plain, India. Environmental Health Perspectives 112: A19–20. Acharyya, S.K. & Shah, B.A. 2004b. Genesis of pandemic arsenic pollution affecting Bengal Basin. National Academy Science Letters 27(5&6), In press. Agrawal, P.K., Basu, A.R. & Poreta, R.J. 2000. Isotope hydrology of groundwater in Bangladesh: Implication for characterization and mitigation of arsenic in groundwater. Preliminary report IAEA-TC project BGD/8/016: 24. Berg, M., Tran, H., Nguyn, T.C., Schertenleib, R. & Giger, W. 2001. Arsenic contamination of groundwater in drinking water in Vietnam: A human health threat. Environmental Science & Technology 35: 2621–2626. Bhattacharya P., Chatterjee, D. Jacks, G. 1997. Occurrence of arsenic-contaminated groundwater in alluvial aquifers from Delta Plain, Eastern India: options for safe drinking water supply. Water Res. Develop. 13: 79–92. Bhattacharya, P., Tandukar, N., Neku, A., Valero, A.A., Mukherjee, A.B. & Jacks, G. 2003. Geogenic arsenic in groundwaters from Terai alluvial plain of Nepal. Jour. de Physique IV France, 107, 173–176. Chakraborti, D., Mukherjee, S.C., Pati, S., Sengupta, M.K., Rahman, M.N., Choudhury, U.K., Lodh, D., Chanda, C.R., Chakraborti, A.K. & Basu, G.K. 2003. Arsenic groundwater contamination in middle Ganga plain, Bihar, India: A future danger? Environmental Health Perspectives 111: 1194–1201. Chitrakar, R.L. & Neku, A. 2001. The scenario of arsenic in Drinking water and Arsenicosis in Nepal. http://groups.yahoo.com/arsenic-source/files/scenario As – Drinking water. Guha, S., Raymahashay, B.C., Banerjee, A., Acharyya, S.K. & Gupta, A. 2004. Collections of depth-specific groundwater samples from arsenic contaminated aquifer in West Bengal, India. (in prep.). Kinnibugh, D.G. & Smedley, P.L. 2001. British Geological Survey Report, WC/00/19. Kumar, G., Khanna, P.C. & Prasad, S. 1996. Sequence stratigraphy of the foredeep and evolution of the IndoGangetic plain, Uttar Pradesh. Proc. Symp. NW Himalaya and Foredeep, Geol. Surv. India, Spec. Pub. 21(2): 173–207. Nickson, R., McArthur, J., Burgess, W., Ahmed, K.M., Ravenscroft, P. & Rahman, M. 1998. Arsenic poisoning of Bangladesh groundwater. Nature 395: 338. Nickson, R., McArthur, J., Ravenscroft, P., Burgess, W. & Ahmed, K.M., 2000. Mechanism of arsenic release to groundwater, Bangladesh and West Bengal. Applied Geochemistry 15(4): 403–413. Pal, T., Mukherjee, P.K., Sengupta, S., Bhattacharyya, A.K. & Shome, S. 2002. Arsenic pollution in groundwater of West Bengal, India – An in sight into the problem by subsurface sediment analysis. Gondwana Research 5: 501–512. Sanyal, S.K. & Naser, S.K.T. 2002. Arsenic contamination of groundwater in West Bengal (India): Build-up in soil-crop systems. Intern. Conf. Water related disasters, Kolkata, Dec’02. E:\Arsenic1\Sanyal-Naser2002-As.doc 4/26/04. Shivanna, K., Sinha, U.K., Sharma, S., Joseph, T.B., Navada, S.V., Roy, A., Talukdar, T., Mehta, B.C. & Ghosh, A.K. 2000. Environmental isotopes analysis for assessing source and mobilisation of arsenic in ground water. Proc. Internat. Workshop on control of arsenic contamination in groundwater. Pub. Health Eng. Dept. Govt. W. Bengal: 72–83. Singh, I.B. 2001. Proxy records of neotectonics, climate changes and anthropogenic activity in the Late Quaternary of Ganga plain. Nat. Symp. Role of Earth Sci. in Integrated Development and Related Societal Issues, Geol. Surv. India, Spec. Pub. 65(1): xxxiii–xlx. Uddin, M.N. & Abdullah, S.K.M. 2003. Quaternary geology and aquifer systems in the Ganges-BrahmaputraMeghna delta complex, Bangladesh. Proc. GEOSAS-IV, Geological Survey of India: 400–416.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Arsenic pollution in groundwater of West Bengal, India: Where we stand? D. Chandrasekharam Department of Earth Sciences, Indian Institute of Technology Bombay, Mumbai, India

ABSTRACT: The arsenic content in groundwater in both shallow and deep aquifers in West Bengal is causing concern. The situation in West Bengal appears to be similar to that reported in Bangladesh, where high arsenic levels in shallow aquifers is related to irrigation practices. Considering the fact that the root zones in paddy fields are enriched in arsenic, it is likely that irrigation practices may establish a continuous chain between high arsenic and low arsenic groundwater. It is high time that large scale canal irrigation is advocated in West Bengal through controlled interlinking of major rivers within Bengal Basin.

1

INTRODUCTION

It is well known that arsenic contamination in groundwater in West Bengal, India and Bangladesh has become a global concern (Das et al. 1996, Bagla & Kaiser 1996, Mallick and Rajagopal 1996, Mandal et al. 1996, Chakraborti et al. 1996, Bhattacharya et al. 1997, Chowdhury et al. 1999, Acharyya et al. 2000, Nickson et al. 2000, Chandrasekharam et al. 2001, McArthur et al. 2001, Ravenscroft et al. 2001). About 44% of total population of West Bengal is suffering from arsenic related diseases like conjunctivitis, melanosis, hyperkeratosis, and hyper pigmentation. In certain areas gangrene in the limb, malignant neoplasm and even skin cancer have also been observed. Here the arsenic content in groundwater varies from 0.05 to 3.7 mg/L, with an average of 0.2 mg/L (Stüben et al. 2003). Although considerable effort was made during the last several years by many organizations to evaluate the scale of contamination and to evolve adequate methods and strategies to reduce the impact of this catastrophe, several questions remained unanswered. Beside assessment of the extent to which the groundwater is contaminated with arsenic, a major thrust should be to find out the source(s) of arsenic in this region. Now it seems certain that the arsenic is of geogenic origin (Chandrasekharam et al. 2001, Stüben et al. 2002). Though several mechanisms of arsenic release into groundwater has been documented from various parts of the world, the source contributing arsenic into the Bengal basin aquifers has yet to be established (Smedley & Kinniburgh 2002, Bhattacharya et al. 2002a). Examples of reducing (e.g. Bengal Basin), oxidizing (Argentina and Chile) and both reducing and oxidizing mechanisms (USA) have been well documented (Smedley & Kinniburgh 2002, Bhattacharya et al. 2002a, Bundschuh et al. 2004). Adsorption and desorptiondissolution by iron and manganese oxides are the commonly recognized processes of arsenic release into the groundwater (Smedley & Kinniburgh 2002, Bhattacharya et al. 2002b, Stüben et al. 2003 and the references therein). Besides naturally controlled redox conditions, West Bengal is experiencing “irrigation controlled” redox conditions, especially in areas where tube-well irrigation is practice. 2

EFFECT OF FLOODS AND IRRIGATION ON THE SHALLOW AQUIFERS

In West Bengal, earlier reports indicate contamination of shallow aquifers (12–15 m) by arsenic; but investigation by several later groups provided evidence of arsenic contamination of deeper 25

aquifers (70–150 m) as well (Chakraborty et al. 1996, Chandrasekharam et al. 2001, Bhattacharya et al. 2002a, Smedley & Kinniburgh 2002, Stüben et al. 2003 and references therein). Besides iron and manganese oxides, experimental investigation indicates a mechanism of arsenic release into the groundwater from soils. At higher soil redox levels (⬃500 mV) arsenic solubility was low while at low redox levels (⫺200 mV) the soluble arsenic content increases by 13 folds as compared to 500 mV (Masscheleyn et al. 1991). Maximum conversion of As(V) to As (III) takes place under redox potential of ⫹100 mV and below. At around ⫹150 mV manganese also gets mobilized into aqueous phase releasing arsenic into the solution. This process has been recognized in the field conditions also around Murshidabad, West Bengal, where dissolution of both iron and manganese oxides are responsible for increasing the arsenic levels in groundwater (Stüben et al. 2003). The most interesting experimental finding, which is very relevant to the conditions

Figure 1.

Drainage map of West Bengal.

26

prevailing in west Bengal quite often, is that reported by Onken & Hossner (1996). Soil solution under flooding conditions recorded maximum concentration of arsenic compared to the soils under non-flooding conditions. Release of soluble arsenic under flooding conditions takes place with in few hours of flooding. The Bengal Basin in general and West Bengal in particular (Fig. 1), invariably gets flooded every year during monsoon periods, which extends from June to December. About 43% of the total area of 89,000 km2 gets flooded during this period (Relief Web 2004 www.reliefweb.int) causing physical damage to more than 30% of population (total population as on 2001 is 80 million). If an analogy is drawn between the experimental work of Onken & Hossner (1996) and the flood conditions that prevail during monsoon period over the Bengal Basin, soil solution appears to be a potential source of arsenic into the shallow aquifers. If only monsoon effect is taken into account, this should be a temporal phenomenon and during non-monsoon seasons arsenic should get adsorbed on to the soils as insoluble species. However, considering the agricultural practices adopted in the state, a large part of the land is always under reducing conditions. As on today, the cultivable land area of West Bengal is about 54,640 km2 and 64% of this area is under irrigation. With ever increasing demand for food, large scale minor irrigation is encouraged and irrigated land area is thus increasing annually. As a result, farmers are able to adopt multiple cropping pattern methods on the same soils and work for twelve months in an year. The system of irrigated farming improved other associated industries like agro-based industries, animal husbandry, fish culture, fertilizer industries and created huge employment for the rural population. According to the Water Investigation and Development Department (WIDD), minor irrigation through tube wells and from rivers was started in 1959 in collaboration with the Exploratory Tube-Well Organization of the government of India. By the year 1976 about 20,000 shallow and deep tube wells were drilled while only 700 lift irrigation from rivers were established. By the end of the year 1985, additional 5634 tube wells were drilled to increase the irrigated land area. On the contrary only 69 river lift irrigation systems were added in this year. Under the Rural Infrastructure Development Fund Plan (RIDF) additional tube wells were added thus increasing an additional 1500 km2 of irrigated land to the existing land. By the year 2001 further 11,150 km2 of land was brought under tube well irrigation scheme and the number of tube wells being used, as per the 2001 census, is about 550,000. Thus, as on date, 47,650 km2 of land is cultivated through irrigation. Successful implementation of minor irrigation schemes changed the shape of rural Bengal. A feel-good factor prevailed in the State under the shadow of a giant arsenic death cloud. Though irrigated farming supported millions of people in West Bengal, this has adversely affected arsenic levels in the shallow aquifers in the State. Relationship between irrigated farming and high arsenic levels in shallow aquifers has been established in a recent study in Bangladesh (Mainuddin 2002). Existence of similar relationship in West Bengal can not be ruled out. It is surprising to know that WIDD which over sees the water availability as well as the quality parameters of groundwater has ignored the quality aspect of water available through shallow and deep tube wells. Had there been a strict quality control on the groundwater exercised by the said department, the magnitude of arsenic poisoning in this state could have been minimized or controlled.

3

EFFECT OF TUBE WELL IRRIGATION ON ARSENIC LEVELS IN GROUNDWATER

In West Bengal, nearly 91% of total cultivated land is occupied by rice production. About 92% of total food produced is rice. Rice crops need standing water column of about a few centimetres above the root zone. Thus, reducing environment prevails just above the root zone. Thus in an year the cultivable and non cultivable lands are under reducing environment, at least for certain time of the year, either due to floods, as described above, or due to rice cultivation. The arsenic content in groundwater drawn through tube wells for irrigation varies from 0.05 to 3.7 ␮g/g (Chandrashekharam et al. 2001, Stüben et al. 2003). Arsenic content in different parts of rice plant (root, stem, nodes, 27

leaves and seed) reveal that the roots absorb maximum amount of arsenic with the content varying from 169 to 178 ␮g/g in the case of rice and 0.3 to 0.7 ␮g/g in the case of wheat while, stem, leaves and grain have relatively low levels of arsenic content (Agarwala 2003, Norra et al. 2004). After harvesting, these roots are ploughed back into the soil during the cultivation of subsequent crop. Thus the arsenic rich groundwater drawn from deep tube wells is pumped back into the shallow aquifers through roots. This is a common practice adopted by all the farmers in India over centuries and this practice will continue in future. If this practice continues, it is likely that the entire shallow (and deeper aquifers) aquifers may get contaminated severely in the very near future.

4

REMEDIAL SOLUTION

The government of West Bengal realized the effect of irrigation on the quality of groundwater and started experimental projects of rainwater harvesting to improve the quality and quantity of groundwater in the state. The Central Groundwater Board (CGWB) has been entrusted with this task for three districts. The Bengal Basin has a well distributed network of surface drainage system and the aquifers are not starved of recharge! How far CGWB will succeed in reducing the incidence of arsenic in groundwater through rainwater harvesting is matter to be examined over a period time. CGWB has reputed hydrogeologists who should have advised the government on this poorly conceived project. This project seems to be like a rain drop in an ocean. Perhaps adopting alternate methods of irrigation may be a viable solution to control arsenic levels in the shallow aquifers. According to The Indian Council of Agricultural Research (ICAR 1992), West Bengal can be divided in to six major agro-ecological sub-regions. The four most important sub-regions are: (1) hot dry sub-humid, (2) hot moist sub-humid (Bengal Basin), (3) hot per-humid, and (4) hot moist sub-humid (Gangetic delta). The hot moist sub-humid regions (e.g. Baharampur in Murshidabad district) have different kind of soil and experiences flooding while hot dry sub-humid regions (e.g. Ahmadpur in Birbhum district) are free from flooding with soil type different from the former region. The most interesting fact is that in the hot moist sub-humid regions agriculture is based on irrigation and crops are grown through-out the year while in the hot dry sub-humid regions rain/canal-fed cultivation is adopted. In both these regions rice is cultivated extensively. The influence of dry sub-humid climate on the release of arsenic from soil solutions is yet to be ascertained. Considering the arsenic levels in the rivers draining the Bengal Basin (⬍1.9 ␮g/L; Bhattacharya et al. 2002a) probability of rice roots accumulating large quantities of arsenic is low, hence multiple crop cultivation may not have adverse effect on the arsenic content in shallow aquifers of such regions. Thus in general, groundwater from deep and shallow aquifers is extensively used in central and eastern regions while surface water irrigation is common along the western region. Though mitigating reducing conditions due to floods is not possible, recycling arsenic groundwater for irrigation can be mitigated. Since rice is cultivated through rain/canal irrigation in large part of West Bengal, attempts should be made to promote similar irrigation culture in the moist sub-humid regions of West Bengal as well. This is possible by creating a net work of surface irrigation canal system through out the Bengal basin. As seen in Figure 1, West Bengal has an excellent network of surface drainage system. The entire Bengal Basin is traversed by several tributaries of the river Ganges. As a first step, on an experimental basis, a network of irrigation canals connecting the drainage may be established in areas marked in Figure 1, and rice cultivation supported by such network of channels should be recommended. As mentioned earlier, the rice roots recorded higher arsenic content in this area. This will provide arsenic free water to the rice crop and prevents accumulation of arsenic in root zone of the plants. Both groundwater in the shallow aquifers and the soil solutions can thus be made free from arsenic. The local government should take initiative to educate the farmers of the advantages of such system for the benefit of the large population. In a long run this will provide safe drinking water as well arsenic free food to the population of West Bengal. 28

REFERENCES Acharyya, S.K., Lahiri, S., Ramahashay, B.C. & Bhowmik, A. 2000. Arsenic toxicity of groundwater in parts of the Bengal Basin in India and Bangladesh: the role of Quaternary stratigraphy and Holocene sea level fluctuation. Environ. Geol. 39: 1127–1137. Agarwala, P. 2003. Arsenic contamination in groundwater, surface water, soils and uptake of arsenic by cultivated crops in Malda district, West Bengal, India. M.Tech. Thesis (unpublished), IIT Bombay, 2003: 109p. Bagla, D.E. & Kaiser, J. 1996. India’s spreading health crisis draws global arsenic experts. Science 274: 174–175. Bhattacharya, P., Frisbie, S.H., Smith, E., Naidu, R., Jacks, G. & Sarkar, B. 2002a. Arsenic in the Environment: A Global Perspective. In: B. Sarkar (ed.): Handbook of Heavy Metals in the Environment, 147–215, Marcell Dekker Inc., New York. Bhattacharya, P., Jacks, G., Ahmed, K.M., Khan, A.A. & Routh, J. 2002b. Arsenic in groundwater of the Bengal Delta Plain aquifers in Bangladesh. Bull. Env. Cont. Toxicology 69: 538–545. Bundschuh, J., Farias, B., Martin, R., Storniolo, A., Bhattacharya, P., Cortes, J., Bonorino, G. & Albouy, R. 2004. Groundwater arsenic in the Chaco-Pampean Plain, Argentina: Case study from Robles County, Santiago del Estero Province. Appl. Geochem. 19(2): 231–243. Chakraborti, D., Das, D., Samanta, B.K., Mandal, B.K., Roy Chowdhury, T., Chanda, C.R., Chowdhury, P.P. & Basu, G.K. 1996. Arsenic in groundwater in six districts of west Bengal, India. Environ. Geochem. Health 18: 5–15. Chowdhury, T.R., Basu, G.K., Mandal, B.K., Biswas, B.K. Samanta, G., Chowdhury, U.K., Chanda, Ch.R., Lodh, D., Roy, S.L., Saha, K.Ch., Roy, S., Kabir, S., Quamruzzaman, Q. & Chakraborti, D. 1999. Arsenic poisoning in Ganges Delta. Nature 401: 545–546. Chandrasekharam, D., Karmakar, J., Berner, Z. & Stüben, D. 2001. Arsenic contamination in groundwater, Murshidabad district, West Bengal. Proceed. Water-Rock Interaction 10, A.Cidu (ed.), 1051–1058, A.A.Balkema, The Netherlands. Das, D., Samanta, G., Mandal, B.K., Chowdhury, T.R., Chanda, Ch.R., Chowdhury, P.P., Basu, G.K. & Chakroborti, D. 1996. Arsenic in six districts of West Bengal, India Environ. Geochem. Health 18: 5–15. ICAR 1992. Indian Council of Agricultural Research Report 48. Masscheleyn, P., Delaune, R.D. & Patrick, Jr. W.H. 1991. Effect of redox potential and pH on arsenic speciation and solubility in a contaminated soil. Environ. Sci. Tech. 25: 1414–1419. Mallick, S. & Rajagopal, N.R. 1996. Groundwater development in the arsenic affected belt of West Bengalsome questions. Curr. Sci. 70: 956–958. Mainuddin, M. 2002. Groundwater irrigation in Bangladesh: ‘tool for poverty alleviation’ or ‘cause of mass poisoning’? Proceed. Sym. Intensive use of Groundwater-Challenges and Opportunities (SINEX), Valencia, Spain, 10–14 December 2002, 152–168. Mandal, K.B., Chowdhury, R., Samanta, G., Basu, K.G., Chowdhury, P.P., Chanda, R.C., Lodh., D., Karan, K.N., Dhar, K.R., Tamili, K.D., Das, D., Saha, K.C. & Chakraborti, D. 1996. Arsenic in groundwater in seven districts of West Bengal-the biggest arsenic calamity in the world. Curr. Sci. 70: 976–986. Norra, S., Aggawala, P., Wagner, F., Berner, Z., Chandrasekharam, D. & Stüben. D. 2004. Impact of irrigation with As rich groundwater on soil and crops: a geochemical case study in Malda District, West Bengal (manuscript in preparation). Nickson, R.T., McArthur, J.M., Ravenscroft, P., Burgess, W.G. & Ahmed, K.M. 2000. Mechanism of arsenic release to groundwater, Bangladesh and West Bengal. App. Geochem. 15: 403–413. McArthur, J.M., Ravenscroft, P., Safiullah, S. & Thirlwall, M. 2001. Arsenic in groundwater: testing pollution mechanisms for sedimentary aquifers in Bangladesh. Water Resour. Res. 37: 109–117. Onken, B.M. & Hossner, L.R. 1996. Determination of Arsenic species in soil solution under flooded conditions. Jour. Soil Sci. Soc. Am. 60: 1385–1392. Ravenscroft, P., McArthur, J.M. & Hoque, B.A. 2001. Geochemical and palaeohydrological controls on pollution of groundwater by arsenic. In W.R.Chappell, C.O. Abernathy and R.L. Calderon (ed) 4th Inter. Conf. on Arsenic exposure and health effects, 112–134, Elsevier Oxford. Relief Web 2004. Situation report – West Bengal floods, 16 Jul 2004. URL: http://www.reliefweb.int (Accessed on 16 July 2004). Smedley, P.L. & Kinniburgh, D.G. 2002. A review of the source, behaviour and distribution of arsenic in natural waters. App. Geochem. 17: 517–568. Stüben, D., Berner, Z., Chandrasekharam, D. & Karmakar, J. 2003. Arsenic pollution in groundwater of West Bengal, India: Geochemical evidences for mobilization of As under reducing conditions. App. Geochem. 18: 1417–1434.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Mineralogical characteristics of the Meghna floodplain sediments and arsenic enrichment in groundwater A.M. Sikder & M.H. Khan Arsenic Research Group [BD], Banani, Dhaka, Bangladesh

M.A. Hasan & K.M. Ahmed Department of Geology, University of Dhaka, Dhaka, Bangladesh

ABSTRACT: The study attempts to investigate the mineralogical composition of the fluvial deposits forming aquifers in order to understand the behavior of the solid phases in relation to arsenic enrichment in groundwater in the ‘Meghna floodplain’ of the Bengal Basin. The major elements in the sediment samples along with its total arsenic content were determined by XRF spectrophotometer. Leachable arsenic in the sediments samples from the zone of oxidation (3–4 m depth) under atmospheric condition at pH 5.5 is insignificant, though the samples contain relatively high amount of total arsenic. Jarosite (KFe3[SO4]2[OH]6) is identified in a number of samples and alunite (KAl3[SO4]2[OH]6) is also observed in the lower part of the sequence. Carbonate minerals are found to exist throughout the entire sequence in the form of siderite (FeCO3), dolomite [CaMg(CO3)2] and rhodocrosite (MnCO3). The present mineralogical and chemical analyses of sub-surface sediment samples of the studied area reveals that oxidation due to the fluctuation of the groundwater level do not contribute significant amount to the arsenic release to groundwater. The strong correlation between TOC content and total arsenic indicates that organic matters might have played an important role in the mobilization of arsenic to the groundwater of the Meghna floodplain of the Bengal Basin.

1

INTRODUCTION

Arsenic enrichment in groundwater at shallow depths in the Ganges-Brahmaputra-Meghna (GBM) delta have been considered as an environmental catastrophe of the current time (Dhar et al. 1997, Bhattacharya et al. 1997, BGS & DPHE 2001, Bhattacharya et al. 2002, Chakraborty et al. 2002, Ahmed et al. 2004). Arsenic occurrences in GBM delta have been reported as a geogenic event whereas the release mechanism into groundwater has been described as reductive dissolution or desportion of arsenic bound to iron oxyhydroxide (Bhattacharya et al. 1997, Nickson et al. 2000, McArthur et al. 2001, Anawar et al. 2002, Smedely & Kinniburgh 2002, Tareq et al. 2003, Akai et al. 2004, Zheng et al. 2004). However, there are some gaps in the precise understanding regarding the state of arsenic in the sediments and the processes that trigger its release and mobilization into groundwater (Mallik & Rajagopal 1996, Acharyya et al. 1999, Yamazaki et al. 2000, Dowling et al. 2002, Harvey et al. 2003). The recent sediments of the basin exhibit a thick succession of fluviatile sediments deposited by the GBM river systems and characterized by complete or truncated cycles of fining upward sequences dominated by coarse to fine sand, silt and clay (Bhattacharya et al. 1997). In general, the GBM sediments consist of about 50–65% quartz, 7–15% feldspar, 7–20% lithic fragments, 5–15% mica with insignificant presence of opaque minerals. The persistence of sulfide group of minerals is quite unusual in the sediments of warm and humid climate, although authigenic pyrite along with corroded iron oxyhydroxides noticed in the fine clastic sediments (Ahmed et al. 2001). 31

Though some mineralogical analyses of the recent sediments of GBM delta have carried out by different investigators, but attempts have not been made to characterize the opaque phases of minerals, in particular. The present study aims to look into the mineralogy of the sediments including characterization of the opaque phase of the minerals and its relation to release mechanism under different geological and geochemical environments.

2

METHOD OF STUDY

An exploratory boring with a modified ‘split spoon’ system was conducted in the ‘Meghna floodplain’ to collect continuous and intact undisturbed sediments samples from the whole sedimentary sequences from 3 to 40 meter covering the modal depths for yielding water through hand tubewells of the present study area (Fig. 1). The mineralogical investigations coupled with chemical analysis of the collected sediment samples were expected to provide useful evidence regarding the state and release mechanism of arsenic. The major elements and the content of total arsenic (As) of the bulk sediment samples have been determined by X-ray fluorescence (XRF) spectrophotometer. A small number of samples with high arsenic content were also analyzed with Instrumental Neutron Activation Analysis (INAA) to validate the results of the XRF analysis. To determine the amount of leachable arsenic from the sediments in the atmospheric condition, 1 gm of sediment sample was dispersed in 100 ml de-ionized water in a beaker and stirred with a magnetic stirrer for 3 hours. The pH of the mixture was measured and maintained to 5.5 by sodium hydroxide as per requirement. Then the homogeneous mixture was filtered through a 45-micron filter using a vacuum pump and the arsenic content was determined with Hydride Generation Atomic

Figure 1.

Map of the study area showing geomorphology and location of the drilled sampling hole.

32

Absorption Spectrophotometer (HG-AAS). The samples from shale intervals were selected to determine the total organic carbon (TOC) content with TOC analyzer using combustion method. Standard procedures were followed for mineralogical analysis of the clay (⬎2 ␮m) fraction with XRD. Special emphasis was given in the preparation of the heavy mineral fractions for XRD analysis to avoid the interference of the major rock forming silicates and the clay minerals. The samples were thoroughly washed and decanted until the clay particles were completely removed. Then selected washed samples were chosen for heavy mineral fraction separation based on the differences in density of the minerals by Heavy Liquids Method (Mullar 1966) for XRD analysis. Samples were prepared in order to obtain free minor constituents (i.e. oxides, hydroxides and sulfides group of minerals) of the sediments for accurate determination of possible minerals containing solid phase arsenic.

3 3.1

RESULTS Lithofacies analysis

The entire lithologic succession was examined from the cored sediment samples (Fig. 2). The zone of oxidation was observed to extend up to a depth of about 6 m from the surface as evidenced by the yellowish-brown color of the sediments. The preservation of few inches thick black patches within the zone of oxidation testifies the presence of high amount of organic matter in the shallower part of the studied sequence. Ferruginous concretions with concentric nature have been observed in the zone of oxidation with decreasing downward tendency in development. These concretions hint at the presence of the microbial activity (i.e. chelation) in the ‘zone of oxidation’. The predominant lithology of the studied sequence was alteration of sand and shale. The light gray sands were fine to medium grained with abundance of mica. The sand beds were characterized by laminations, cross-laminations with occasional trough cross-laminations and clay drapes. A medium to coarse-grained micaceous yellowish brown thickly bedded sand sequence was encountered between 21 and 30 m. Clay drapes were also observed within this sandy sequence.

Figure 2. Depth profiles of total arsenic content, leachable arsenic at pH 5.5, total organic carbon (TOC) in the studied sedimentary sequence.

33

Thick layers of dark gray shale with sand lenticles were observed in the studied sequence at number of depths. A 5 cm thick peat layer was encountered at a depth of 38 m. The sediments were also characterized by the occurrence of decomposed and partially decomposed plant debris. Organic matter was also found to concentrate within the dark the laminations. The studied section exhibited cycles of fining upward sequences (Fig. 2) which can be interpreted as floodplain deposit with strong tidal influence based on the presence of clay drapes and sand lenticels through out the sequence. The coarser sandy horizons in the middle and lower part of the sequence can also be interpreted as the product of high-energy deposition based on the textural and sedimentological characteristics (i.e. flood surge). 3.2

TOC of the sediments

TOC content of sediments were considered to play a significant role in the arsenic enrichment processes. Besides the occurrence of patches of organic matter in the zone of oxidation and the peat layer in the lower part of the sequence, the studied sediments contain relatively high amount of TOC in the range of 0.3 to 1.01%. The maximum concentrations of TOC were found at the peat layer where it reaches to 45.7% (Table 1). Concentrations of organic matter in different layer were even visible with naked eye. 3.3

Major elements

Oxides of the major elements of the selected samples were determined to find out the possible link of arsenic and other elements in the sedimentary sequence (Table 1). Determination of the total

Table 1.

Distribution of the major elements, TOC and As in the sediments.

Depth (m)

SiO2 (%)

Al2O3

Fe2O3

TiO2

MgO

CaO

Na2O

K2O

MnO

P2O5

As* (mg/L)

TOC (%)

4.3 6.1 8.3 10.1 10.1 13.5 17.8 17.8 20.5 21.5 21.5 26.7 26.7 27.6 27.6 27.9 27.9 33.1 35.3 37.1 37.1 38.0 39.3

62.58 60.81 66.15 73.24 61.93 65.33 69.42 48.74 70.03 62.15 68.15 60.61 73.11 66.73 72.15 60.66 56.92 55.68 56.18 69.75 58.71 15.05 66.50

16.01 16.09 14.40 10.92 16.16 11.37 12.37 14.86 11.95 15.96 12.77 14.95 11.23 12.43 11.46 16.81 18.31 13.42 18.39 11.63 16.81 0.54 10.34

6.53 6.79 6.61 3.91 7.01 7.76 5.30 14.85 4.88 6.96 5.25 7.17 3.91 6.60 5.00 7.18 7.73 7.31 7.83 5.11 7.17 0.49 5.35

0.73 0.79 0.71 0.52 0.80 1.26 0.56 1.13 0.67 0.81 0.61 0.83 0.52 0.74 0.75 0.79 0.68 0.82 0.81 0.66 0.80 0.05 0.68

2.28 2.70 2.47 1.50 2.73 2.71 2.00 5.36 1.74 2.65 2.02 2.87 1.38 2.40 1.64 2.72 2.83 2.53 2.92 1.69 2.84 0.17 1.94

1.17 1.44 1.71 2.28 1.84 5.76 1.92 1.14 2.43 1.73 2.02 1.97 2.06 1.24 2.46 1.28 1.17 0.69 0.99 2.22 1.32 0.22 1.36

1.52 1.45 1.69 1.93 1.59 1.60 1.88 1.05 1.90 1.56 1.97 1.66 1.98 1.44 1.81 1.32 1.31 1.12 1.20 1.86 1.42 0.21 1.63

3.13 3.19 2.96 2.22 3.06 1.44 2.65 3.26 2.39 3.00 2.69 3.00 2.41 1.99 2.19 3.05 3.26 3.26 3.38 2.40 3.27 0.12 2.64

0.10 0.12 0.10 0.07 0.11 0.18 0.08 0.22 0.09 0.12 0.08 0.12 0.06 0.13 0.09 0.12 0.14 0.10 0.11 0.09 0.12 ⬍0.01 0.07

0.12 0.10 0.10 0.12 0.11 0.39 0.12 0.15 0.14 0.12 0.12 0.12 0.10 0.11 0.18 0.11 0.12 0.10 0.10 0.14 0.10 0.03 0.10

46 50 22 11 51 7 18 80 15 55 21 39 11 69 14 70 61 109 83 13 59 175 37

0.34 0.74 0.81 0.38 0.65 0.33 0.46 1.49 0.44 0.58 0.49 0.72 0.39 0.57 0.37 0.60 0.69 1.01 0.70 n.a** 0.73 45.7 0.48

* Concentration of leachable arsenic from the sediments; ** Not analyzed.

34

arsenic content of the sediments and the understanding of its release mechanism were the main focus of the study. The data were plotted against depth (Fig. 3) to see whether there is any correlation of distribution between the major elements and the arsenic content in the studied sedimentary sequence. 3.4

Determination of leachable arsenic

The determination of the amount of leachable arsenic from sediments of different depths is very important to understand the release mechanism from sediments to water. Sediments from different depths were selected to determine the amount of arsenic released at a controlled condition of pH 5.5 under the atmospheric condition. In selection of the samples emphasis was given to ‘zone of oxidation’ (up to a depth of about 6 m) and the zone where most of the hand tube wells in the area were screened (i.e. most abstracted shallow aquifer). The results of leaching tests are presented in Table 2. The depth profiles of leachable arsenic along with lithology, total arsenic and total organic carbon are shown in Figure 2. 3.5

Mineralogical analyses

The present study was primarily designed to find out the mineralogical aspects of the sediments in order to understand the behavior of the solid phase minerals in relation to the arsenic enrichment in the groundwater at shallow depths. 3.5.1 Normative calculation of mineralogical composition Normative calculations were carried out to determine the modal mineralogical composition of the sediment samples using a computer program ‘SEDNORM’ (Cohen & Ward 1991). The calculated composition of minerals is plotted against depth to develop a theoretical basis to formulate the strategy of further analysis. 3.5.2 Clay mineralogy The relative abundance of the clay minerals in the clay samples were determined based on the relative intensity of the strongest peak of the constituent minerals in XRD traces (see discussions below).

0.00 1.00 2.00 3.00 4.00

0.00

0

0

0

0

-5

-5

-5

-5

-10

-10

-10

-10

-15

-15

-15

-15

-20

-20

-20

-20

-25

-25

-25

-25

-30

-30

-30

-30

-35

-35

-35

-35

-40

-40

-40

-40

Depth (meter)

0

0 5 10 15 20 25 30

50 100 150 200

-45

Figure 3.

-45

-45

-45 Total As (mg/kg)

Fe2O3%

5.00

MnO2%

CaO%

Depth profile of elements showing the relationship with total arsenic content of sediments.

35

10.00

Table 2.

Results of leaching test of the sediments samples.

Depth (m)

Arsenic in sediments (mg/kg)

Arsenic in leachate (mg/L)

3.1 3.4 4.0 6.4 8.3 10.4 16.3 17.2 23.0 23.9 24.5 24.8 25.1 25.8 28.2 33.7 35.9 38.0 39.2

60* 84* 19.8* 49.7* 2.48 62.10* 2.35 0.72 2.02 2.78 3.50 3.54 3.64 0.74 63.1* 109 83 175 37

Leaching does not occur Leaching does not occur Leaching does not occur 0.071 0.252 0.103 0.112 0.14 0.124 0.069 0.099 0.216 Leaching does not occur Leaching does not occur Leaching does not occur 0.085 0.12 Leaching does not occur Leaching does not occur

* INAA. Table 3.

Results of XRD analysis of heavy mineral fractions (specific gravity ⬎ 2.9).

Sample no.

Depth (m)

Weight (%)

Significant minerals

1 2 3 4 5 6 7 8 9 10 11 12

4.6 5.2 9.2 12.9 15.0 19.0 20.8 23.0 24.2 25.1 32.8 35.9

2.11 1.93 4.26 3.15 5.68 5.45 3.27 4.27 3.76 4.99 3.32 1.68

Jarosite Ankerite, Magnetite, Dolomite Pyrolusite, Magnetite, Amphiboles, Kaolinite Pyrolusite, Magnetite, Amphiboles, Siderite Pyrolusite, Amphiboles Magnetite, Amphiboles, Kaolinite Magnetite, Amphiboles, Kaolinite Jarosite, Pyrolusite, Amphiboles, Dolomite Pyrolusite. Magnetite, Amphiboles Alunite, Pyrolusite, Amphiboles Jarosite, Amphiboles, Kaolinite Pyrolusite, Amphiboles

3.5.3 XRD of heavy mineral fraction The environmental sensitive minerals were identified by removing the major constituents of the sediments (i.e. clay and silicates group of minerals) in order to avoid their interference in the analysis. The main concern of the analyses was to find the forms in which metals and metalloids move through the hydrosphere. Thus attention was given in the identification of minerals that do not show up very often in routine mineralogical analysis. The significant minerals identified from the XRD traces are listed in the Table 3. 4

DISCUSSION AND CONCLUSIONS

The plots of relevant oxides (Fig. 3) revealed that among the major elements, arsenic showed good correlation only with iron oxide and manganese oxide (Fe2O3 and MnO). The total arsenic content 36

was moderately high in the samples of shallower depth including those from the ‘zone of oxidation’. Highest arsenic content was observed (109 mg/kg) in the sample from 33 m corresponding with a shale sequence. Potassium oxide (K2O) and magnesium oxide (MgO) also showed well conformity with the total arsenic content of the sediments (Table 1). These conformities were strong in the deeper parts than the shallower part of the sequence. This was probably due to the weathering effects in the shallower depth. But the sequences in the deeper parts, not affected by the weathering process, showed a good correlation among potassium (K), magnesium (Mg) and arsenic (As) and that was probably due to the distribution of those elements from the source of sediments. The most common K bearing minerals in sediments is muscovite (K2 Al4 [Si6Al2O20] (OH)4]) whereas the source of Mg is generally the ferromagnesian minerals (i.e. amphiboles and biotite) and to a lesser extent magnesite (MgCO3). Manganese oxide (MnO), titanium oxide (TiO2), sodium oxide (Na2O) and calcium oxide (CaO) do not show any correlation in the distribution with Arsenic content of the sequence (Table 1 and Fig. 3). The samples covering the ‘zone of oxidation’ ranging in depth from 3 to 4 m did not leach any arsenic in the solution at pH 5.5, although XRF analysis of bulk samples showed the presence relatively high amount of total arsenic in the same samples. Significant amount of leaching of arsenic occurred from the coarser sequences ranging in depth from 8 to 17 m that corresponded with shallow aquifer of the study area. The modal and mean depth of depletion occurred at a depth of 25 m. The shale samples of 25 to 28 and 38 to 39 m depths also did not leach any arsenic in the laboratory experiments at pH 5.5 although the total arsenic contents of the corresponding samples were also very high. The clay minerals (2–4 ␮m in size) are considered as the most sensitive indicator of the geochemical environment and weathering processes. Clay minerals can also bind elements in their interlayer and the binding capacity of clays directly depends on the development and presence of certain type of clay minerals (e.g. halloysite and mixed-layer clays) (Paquet & Clauer 1997). The formation of the clay minerals in the shallower depth is scanty, which corresponds with high activity zone of weathering. But the studied sediment samples are found rich in mica. The presence of mica in the clay fraction is also strikingly very high. The increase of the kaolinite content with depth indicates the increase of acidity and decrease of dissolved silica in the pore fluid. Absence of smectite throughout the sequence and weak presence of non-swelling illite-smectite as revealed from the glycol solvated XRD traces of the samples indicate weak acidic and low silica groundwater conditions. The presence of kaolinite throughout the sequence with few occurrences of illite-smectite mixed layer clays hints that the studied sequence suffered less pervasive hydrolysis. Kaolinite is usually observed in depths where acidity is high and dissolved silica is low. So the increase in the amount of kaolinite with depth probably is due to the increase of acidity of the pore fluids with depth (Fig. 4). Jarosite was identified in more than one sample. Jarosite (KFe3[SO4]2[OH]6) is a rather widely distributed mineral in oxidized zone and frequently occurs as a weathering product of pyrite (FeS2) in certain sediments. Jarosite is not very stable in water and yields iron hydroxides on hydrolysis and therefore supposed to be a rare mineral in sediments of warm humid climate (Batekhtin 1981). Alunite (KAl3[SO4]2[OH]6) was also observed in the lower part of the studied sequence. Similar occurrences of Alunite in little deeper depth had also been reported from other areas of the country (Fig. 5). Magnetite (Fe2O3) is observed in most of the samples except from those of the ‘oxidizing zone’ and that implies, magnetite were present in the entire sequence probably right form the source (Fig. 6). Pyrolusite (MnO2) was also identified in most of the samples as the end product of alteration in the zone of low degree of oxidation (Batekhtin 1981, Deer et al. 1967). The expected mineral association of pyrolusite in sedimentary environment is goethite, hematite, lepidocrosite, magnetite, calcite and quartz. Goethite may also be present in the sand sequence but probably not identified in XRD traces due to the low resolution of the tool and sensitivity of 37

Figure 4. XRD trace of oriented clay fraction (⬎2 ␮m) showing the variations of the development of clay minerals with depth.

Figure 5.

XRD traces showing the typical peaks of jarosite and alunite identified in the studied sediments.

chosen radiation. A mineral is possible to identify in XRD if it forms at least 5% of the bulk composition. Also the source of radiation (CuK␣) used in the present analysis was not suitable for the identification of the iron oxide minerals (Batekhtin 1981, Deer et al. 1967). Carbonate minerals were observed throughout the sequence in the form of ankerite (Ca, Fe, Mg, Mn (CO3)2), siderite (FeCO3), dolomite (CaMg(CO3)2) and rhodocrosite (MnCO3). The intensity of occurrence of the carbonates is higher in the shallower depth. The presence of carbonates in the zone of oxidation could have an intimate relationship with chelation of the metals due to the microbial activity, which also correspond with the observation of the high arsenic content in samples 38

Figure 6.

XRD trace showing the major peaks of magnetite and pyrolusite in the studied sediments.

from the shallower depth but without any leaching of arsenic. Chelated metals tend to be very tightly bound thus more insoluble than the other solid phases of metals (Paquet & Clauer 1997). The depth profile of TOC and the arsenic content exhibits quite well conformity. Especially the arsenic content of the peat sample (145 mg/kg) is very significant and provides an important clue for the future investigations. Almost all the areas of the Meghna floodplain go under water during the monsoon and water remains over the land for at least for 3 months. Strict demarcation of the aerobic (oxidizing) and anaerobic (reducing) zoning appeared incompatible at the studied area to some extent, as the oxidizing zone change to reducing zone for a certain period in every year. Although it appears from the present chemical and mineralogical analysis of sub-surface sediments samples of the study area that oxidation due to the fluctuation of the groundwater level do not contribute any significant amount of arsenic to groundwater. In consideration of the present geomorphic and groundwater scenario the detection of jarosite and alunite in the recent sedimentary sequence of GBM is very unusual. But the presence of peat, iron stained sand and dispersed organic matter throughout the studied sequence supports that the geochemical environment favorable for the formation of jarosite and alunite might have prevailed in the Meghna Plain due to frequent and strong tidal incursions in a costal and lagoonal setting (Breemen 1975, Breemen & Pons 1978, Kazutak 1992). Weak acidic groundwater condition with less pervasive hydrolysis in the unweathered section of the studied sequence as revealed from the clay mineralogical analysis support to some extent the presence of Jarosite in the ‘zone of oxidation’. However the explanation of the presence of Jarosite in the reducing environment in deeper part of the sequence other than the preservation of the weathering products of geological past due to rapid subsidence demands further investigations. Lack of advanced analytical facilities limited the study to a large extent. Detailed sediment logical and pore water investigations are essential to draw definite conclusions about the solid phase arsenic and subsequent mobilization to groundwater. REFERENCES Acharyya, S.K., Chakraborti, P., Lahiri, S., Raymahashay, B.C., Guha, S. & Bhowmik, A. 1999. Arsenic posing in the Ganges Delta, Nature 401: 545. Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A.H., Imam, M.B., Khan, A.A. & Sracek, O. 2004. Arsenic contamination in groundwater of alluvial aquifers in Bangladesh: An overview. Applied Geochemistry 19(2): 181–200.

39

Ahmed, K.M., Imam, B., Akhter, S.H., Hasan, M.A. & Khan, A.A. (2001). Sedimentology and mineralogy of arsenic contaminated aquifers in the Bengal Delta of Bangladesh. In Jacks, G., Bhattacharya, P. & Khan, A.A. (Eds.): Groundwater Arsenic Contamination in the Bengal Delta Plain of Bnagladesh, Proceedings of the KTH-Dhaka University Seminar, University of Dhaka, Bangladesh. KTH Special Publication, TRITAAMI Report 3084. Akai, J., Izumi, K., Fukuhara, H., Masuda, H., Nakano, S., Yoshimura, T., Ohfuji, H., Anawar, M.H. & Akai, K. 2004. Mineralogical and geomicrobiological investigations of groundwater arsenic contamination in Bangladesh. Applied Geochemistry 19(2): 215–230. Anawar, H.M., Akai, J., Komaki, K., Terao, H., Yoshioka, T., Ishizuka, T., Safiullah, S. & Kato, K. 2003. Geochemical occurrence of arsenic in groundwater of Bangladesh: sources and mobilization processes. J. Geochem. Explor. 77: 109–131. BGS & DPHE 2001. Arsenic Contamination of Groundwater in Bangladesh. Vol. 2 Final Report, BGS Technical Report WC/00/19. Bhattacharya, P., Chatterjee, D. & Jacks, G. 1997. Occurrence of Arsenic-contaminated Groundwater in Alluvial Aquifers from Delta Plains, Eastern India: Options for Safe Drinking Water Supply. Water Resources Development 13(1): 79–92. Bhattacharya, P., Jacks, G., Ahmed, K.M., Khan, A.A. & Routh, J. 2002. Arsenic in groundwater of the Bengal Delta Plain aquifers in Bangladesh. Bull. Env. Cont. Toxicol. 69: 538–545. Batekhtin, A. 1981. A Course of Mineralogy, Translation from the Russian, Peace Publishers, Moscow. Breemen, Van 1975. Acidification and deacidification of costal plain soils. Soil. Sci. Soc. Amer. Proc. V. 39, pp. 1153–57. Breemen, Van & Pons 1978. Acid Sulfate Soil and Rice. In: Soil and Rice. IRRI, Philipines, pp. 139–761. Chakraborty, D., Rahman, M.M., Paul, K., Chowdhury, U.K., Sengupta, M.K., Lodh, D., Chanda, C.R., Saha, K.C. & Mukherjee, S.C. 2002. Arsenic Calamity in the Indian subcontinent What lessons have been learned? Talanta 58: 3–22. Cohen, D. & Ward, C.R. 1991. SEDNORM; a program to calculate a normative mineralogy for sedimentary rocks based on chemical analyses. Computers and Geosciences 17(9): 1235–1253. Deer, W.A., Howie, R.A. & Zussman, J. 1967. Rock Forming Minerals: Non-Silicates, 5th Impression, Longman, London. Dhar, R.K., Biswas, B.K., Samanta, G., Mandal, B.K., Chakraborti, D., Roy, S., Jafar, A., Islam, A., Ara, G., Kabir, S., Khan, A.W., Ahmed, S.A. & Hadi, S.A. 1997. Groundwater arsenic calamity in Bangladesh. Current Science 73(1): 48–59. Dowling, C.B., Poreda, R.J., Basu, A.R. & Peters, S.L. 2002. Geochemical study of arsenic release mechanisms in the Bengal Basin groundwater. Wat. Resour. Res. 38: 1173–1190. Harvey, C.F., Swartz, C.H., Badruzzaman, A.B.M., Keon-Blute, N., Yu, W., Ali, M.A., Jay, J., Beckie, R., Nieden, V., Brabander, D., Oates, P.M., Ashfaque, K.N., Islam, S., Hemond, H.F. & Ahmed, M.F. 2002. Arsenic mobility and groundwater extraction in Bangladesh. Science 298: 1602–1606. Kazutak, K. 1978. Coastal low land ecosystem in southern Thailand, Malaysia. Ph.D. thesis Kyto University, Japan, pp. 416. Mallik, S. & Rajagopal, N.R. 1996. Groundwater development in the arsenic-affected alluvial belt of West Bengal- Some questions. Current Science 70(11): 956–958. McArthur, J.,M., Ravenscroft, P., Safiullah, S. & Thirlwall, M.F. 2001. Arsenic in groundwater: testing pollution mechanisms for sedimentary aquifers in Bangladesh. Water Resources Research 37(1): 109–117. Mullar, G. 1966. Methods in Sedimentary Petrology. Hafner Publishing Company, New York/London. Nickson, R.T., McArthur, J.M., Ravenscroft, P., Burgess, W.G. & Ahmed, K.M. 2000. Mechanism of arsenic release to groundwater, Bangladesh and West Bengal. Appl. Geochem. 15(4): 403–413. Paquet, H. & Clauer, N. 1997. Soil and Sediments: Mineralogy and Geochemistry. Springer-Verlag, Heidelberg and New York. Smedley, P. & Kinniburgh, D.G. 2002. A review of the sources, behaviour, and distribution of arsenic in natural waters. Applied Geochemistry 17: 517–568. Tareq, S.M., Safiullah, S., Anawar, H.M., Rahman, M.M. & Ishizuka, T. 2003. Arsenic pollution in groundwater: a self-organizing complex geochemical process in the deltaic environment, Bangladesh. Sci. Tot. Environ. 313: 213–226. Yamazaki, S., Ishiga, H., Dozen, K., Higashi, N., Ahmed, F., Sampei, Y., Rahman, M.H. & Islam, M.B. 2000. Geochemical compositions of sediments of Ganges delta of Bangladesh – arsenic release from the peat? Chikyu Kagaku (Earth Science) 54: 81–93. (In Japanese with English abstract). Zheng, Y., Stute, M., van Geen, A., Gavrieli, I., Dhar, R., Simpson, H.J., Schlosser, P. & Ahmed, K.M. 2004. Redox Control of Arsenic Mobilization in Bangladesh Groundwater. Applied Geochemistry 19(3): 201–214.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Naturally occurring arsenic in groundwater of Terai region in Nepal and mitigation options Nirmal Tandukar Department of Water Supply & Sewerage (DWSS), Kathmandu, Nepal

Prosun Bhattacharya, Gunnar Jacks & Antonio A. Valero Groundwater Arsenic Research Group, Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden

ABSTRACT: Natural arsenic (As) was detected in groundwaters in the Terai Alluvial Plain (TAP) in southern Nepal in the year 1999. By the end of March 2004, about 245,000 wells have been tested for As, out of which about 3% samples are found to have exceeded Interim Nepalese Standard of 50 ␮g/L. From the detail study conducted in hotspot district Nawalparasi, natural rocks are thought to be the sources of As that are leached mainly due to the weathering of As bearing minerals from the Himalayas towards the northern Nepal. In this paper, the chemistry of groundwater from highly arsenic affected Nawalparasi district in the central part of the TAP in southern Nepal has been presented. TAP groundwaters are found to be predominantly of reducing character with low ⫺ SO42⫺ and NO⫺ 3 , but high HCO3 concentrations. Total arsenic (Astot) concentration in groundwater varied from 1.7 ␮g/L to as high as 404 ␮g/L. As(III) species is found to be predominant along with elevated levels of dissolved Fe and Mn. The correlation between DOC, HCO⫺ 3 , Fetot and Astot strongly supports the hypothesis of reductive dissolution of Fe-oxyhydroxides as the main mechanism of mobilization of As in groundwater in TAP. Blanket testing by As-field test kits is the easiest way to find out As free sources nearby for tubewell switching. In the absence of As-free source, the only available option is the treatment of water either at the point of entry or at the point of use to meet the drinking water standard. DWSS in collaboration with UNICEF and WHO is conducting blanket testing of As in 10 Terai districts. Based on the blanket test result, As treatment methods such as 3-gagri filter, arsenic biosand filter etc., which are simple, effective, affordable and socially acceptable will be provided as a short-term option to the affected communities in hotspot areas.

1

INTRODUCTION

Extraction of groundwater in Nepal started during International Water Supply and Sanitation Decade about 30 years back by installing tubewells to provide microbially safe water. However the presence of natural arsenic (As) in groundwater has become a global problem especially in Asian continent. Arsenics was detected very recently in groundwaters in Terai Alluvial Plain (TAP) in southern Nepal (Tandukar 2000, Tandukar et al. 2001, Valero 2002, Bhattacharya et al. 2003). The population of Nepal is about 23.4 million among which about 47% live in the 20 Terai districts and about 90% of these people are dependent on groundwater for drinking and other purposes (Fig. 1). From the study conducted so far each Terai district has of the order of about 25,000 tubewells, out of this about 85% tubewells are privately owned. By the end of March 2004, about 245,000 tubewells have been tested for As, out of which about 3% samples are found to have exceeded Interim Nepalese Standard of 50 ␮g/L (NSCA 2001). This paper presents the chemistry of arsenic-rich groundwaters of Nawalparasi district in the central part of the TAP in southern Nepal. 41

Figure 1. Map of Nepal showing the districts (marked with stars) with elevated As concentration in groundwater (based on Tandukar et al. 2001).

Sunawal

Somnath

Sunawal 27.6

Clay

Choti Pratappur

Khairani

Badera Chowk 27.58

Sand

Badera

Gravel

27.56

Sand-silty clay Silty sand-clay

Sukrauli

Gravel mixed with silt Clay

Sand-silty clay

Sand-silty clay

30

27.52

Bairihawa

Hakui

27.5

27.48 Kushma 27.46

Coarse sand with gravel

27.44

0

2 km 83.64

Gravel

Basahi

Kasipur

Tilakpur Magarmudha Radhanagar Thulo Kumuwar Imlitole Ahirauli 27.54 SukrauliGhodpali Manari Pokharapali Kumuwar Baikunthapur

Sand

Sand-silty clay Gravel

Swathi

4 83.66

Rampurwa 83.68

83.7

83.72

83.74

Rampurwa

Clay

Clay Sand with gravel Gravel

Silty sand-clay

Sand-silty clay 0 m

Gravel Clay

Figure 2. Schematic lithology of the selected boreholes in Nawalparasi. Sampling locations are shown in the inset map.

42

2

LOCATION AND GEOLOGY OF THE STUDY AREA

TAP is the northern extension of Indo-Gangetic Plain. It consists of 20 districts including Nawalparasi with a population of about 11.5 million. It has an average width of 30–40 km and altitude ranging from 60–310 m above mean sea level (Anonymous, 2003). Nawalparasi district lies in the Western Development Region of Nepal occupying the total area of 2162 km2 and has a population of about 0.56 million. The length of the highway linking the capital Kathmandu with the Ramgram Municipality (Parasi) is about 260 km. The average rainfall in the region is ca. 2381 mm (1997–2001). In general TAP has geology, which is similar to Bengal Delta Plain (BDP) and is represented by thick clastic sequence of Holocene age comprising inter-locked alluvial deposit of the wider Ganges Plain (Bhattacharya, 2002; GWRDB-UNDP, 1989). The general flow of groundwater is from North to South. The lithology of the aquifers (Fig. 2) shows the sequence of gravel and sand-silty clay-clay sequence, which has been exploited for groundwater abstraction. 3

MATERIALS AND METHODS

27 private and public tubewells extending to a depth of 7.6 to 54.9 m were sampled in the western part of Nawalparasi district. The well locations were marked using Global Positioning System (GPS). Water samples were collected following the procedures of Bhattacharya et al. (2002) which included: (i) filtered (using 0.45 ␮m filters) for the analysis of major anions (ii) filtered and acidified with supra pure HNO3 for the analysis of cations and trace elements including As. Speciation of As(III) was carried out in the field using disposable cartridges following the method as described by Meng & Wang (1998) and Meng et al. (2001). Major cations and trace elements including As were analyzed by Varian Vista-PRO Simultaneous ICP-OES equipped with a SPS-5 autosampler. Major anions like Cl⫺, SO42⫺ were analyzed with a 3⫺ Dionex DX-120 ion chromatograph using an IonPac As14 column. NO⫺ 3 and PO4 were analyzed with Tecator Aquatec 5400 spectrophotometer using the wavelength of 540 nm and 690 nm respectively. 4

GROUNDWATER CHEMISTRY

Groundwater samples were near neutral to alkaline with the pH in the range between 6.1 to 8.1. Field measured redox potential varied in the range between ⫺0.197 to ⫺0.105 V, which suggest fairly reduced condition in the aquifer. The concentration of SO42⫺ (0–133 mg/L) and NO3⫺ (up to 10.8 mg/L) were low. Total arsenic (Astot) concentration were found in the range 1.7–404 ␮g/L with 79–99.9% as As(III) species. Concentration of total Fe (Fetot) and Mn ranged between 0.11–16.4 mg/L and 0.01–1.95 mg/L respectively. Levels of DOC ranged between 15.2–31.9 mg/L (Table 1). The groundwaters were predominantly of Ca-Mg-HCO3 type with HCO⫺ 3 as the principal anion with concentration ranging between 332–549 mg/L (Fig. 3). Total iron (Fetot) concentrations in these groundwaters were positively correlated with Astot (R2 ⫽ 0.59) and DOC (R2 ⫽ 0.56) especially at depths below 20 m (Fig. 4). A positive correlation was observed between Astotal and HCO3 (R2 ⫽ 0.54). Likewise, a strong correlation was observed between DOC and HCO3 (R2 ⫽ 0.68). A strong correlation was noted between As(III) and NH⫹ 4 (R2 ⫽ 0.89) and DOC (R2 ⫽ 0.79). The concentration of As exceeding the Interim Nepalese standard of 50 ␮g/L was found in the depth range of 7–35 m. 5

DISCUSSION

The hydrogeochemical data for groundwater of the TAP aquifer suggest a predominantly reducing 2⫺ ⫺ character with high HCO⫺ 3 , low SO4 , and NO3 concentrations. This is further supported by the 43

44

27.535 27.544 27.550 27.553 27.574 27.592 27.591 27.613 27.607 27.597 27.577 27.559

N-15 N-16 N-17 N-18 N-20 N-21 N-22 N-23 N-24 N-25 N-26 N-27

83.715 83.723 83.726 83.742 83.734 83.704 83.688 83.651 83.646 83.649 83.661 83.670

(deg. E) 83.711 83.689 83.674 83.624 83.627 83.661 83.664 83.695 83.691 83.688 83.670 83.676 83.668

19.8 7.6 29.0 29.0 43.9 35.1 10.7 27.4 54.9 10.7 10.7 10.7

(m) 24.4 13.0 13.7 15.2 16.8 17.4 16.8 18.3 19.8 16.8 16.8 19.8 245.0

Note: bdl – below detection limit.

(deg. N) 27.438 27.474 27.520 27.538 27.522 27.534 27.537 27.542 27.537 27.528 27.539 27.544 27.545

N-1 N-2 N-3 N-4 N-5 N-6 N-7 N-8 N-10 N-11 N-12 N-13 N-14

6.61 6.24 6.81 6.74 6.81 6.89 6.86 7.16 7.12 6.80 7.37 7.75

7.46 6.88 7.36 6.85 7.29 6.65 6.09 6.62 7.40 6.50 6.49 6.73 6.97 10.5 261.0 1.1 45.6 0.7 0.4 25.1 0.9 0.6 33.4 76.6 1.8

0.4 bdl bdl bdl bdl bdl bdl 1.9 0.1 bdl 2.6 0.1

508 502 453 504 407 355 460 407 381 459 478 353

⫺0.147 ⫺0.169 ⫺0.168 ⫺0.197 ⫺0.168 ⫺0.168 ⫺0.178 ⫺0.154 ⫺0.144 ⫺0.188 ⫺0.131 ⫺0.175

NO⫺ 3

(mg/L) (mg/L) (mg/L) 369 3.74 0.0 521 21.5 5.4 332 0.8 bdl 427 32.0 bdl 440 1.0 0.1 476 3.4 1.2 871 42.0 0.4 505 9.2 bdl 428 11.0 1.3 465 1.4 0.1 525 18.0 0.2 408 0.6 bdl 444 1.9 bdl

HCO⫺ Cl⫺ 3

(V) ⫺0.179 ⫺0.120 ⫺0.149 ⫺0.159 ⫺0.105 ⫺0.174 ⫺0.119 ⫺0.136 ⫺0.129 ⫺0.141 ⫺0.131 ⫺0.131 ⫺0.125

Eh

3.1 133.0 bdl 6.2 1.4 bdl bdl bdl 1.3 0.3 27.6 0.1

(mg/L) 0.1 73.7 bdl 0.4 bdl 2.3 59.6 bdl 2.5 2.8 0.1 bdl 0.5

SO42⫺

K⫹ Mg2⫹

34.5 77.7 80.5 64.1 70.7 39.5 11.1 36.8 23.5 15.4 37.1 24.2

1.8 6.2 0.9 1.3 1.0 1.1 2.1 1.4 2.0 2.0 1.7 0.8

25.7 42.9 19.9 34.5 17.9 18.4 39.7 20.7 21.6 34.3 12.9 15.4

(mg/L) (mg/L) (mg/L) 18.5 4.2 18.5 51.7 87.6 37.9 11.8 1.1 14.5 38.1 1.6 29.1 24.0 1.4 30.5 57.8 2.0 28.8 30.6 155.5 49.4 60.7 2.2 41.1 53.3 1.7 32.1 36.4 2.1 21.8 53.6 1.6 29.4 41.0 1.2 20.6 91.1 1.3 10.2

Na⫹

Geochemical characteristics of the groundwater samples from Nawalparasi district, Nepal.

Sample Latitude Longitude Depth pH

Table 1.

117.6 226.4 49.2 90.8 51.8 70.8 116.4 84.8 86.4 116.5 172.1 83.8

(mg/L) 98.7 145.6 93.5 89.8 91.9 68.0 147.0 66.2 69.0 102.4 96.9 80.3 35.5

Ca2⫹

22.7 18.0 16.4 19.5 16.0 15.2 17.6 17.7 15.3 19.5 18.6 16.8

(mg/L) 31.9 19.4 28.1 28.2 25.1 27.6 25.6 25.5 25.5 22.7 21.2 21.1 20.2

DOC

2.5 1.9 0.3 0.9 0.5 0.6 2.1 1.2 0.1 2.9 0.0 0.7

(mg/L) 0.7 0.0 0.2 1.6 0.9 1.7 0.0 2.4 0.6 3.2 1.4 0.7 0.1

NH⫹ 4

153.9 81.9 118.0 177.8 33.9 100.1 314.2 91.8 11.1 67.6 1.7 75.3

(␮g/L) 78.0 2.5 20.0 265.4 34.9 272.3 3.1 409.4 385.1 120.4 120.3 65.7 19.0

Astot

151.4 81.3 108.1 170.2 33.2 92.2 313.9 87.9 10.2 62.2 1.4 69.0

(␮g/L) 77.9 2.0 19.0 262.5 33.7 270.6 3.0 397.8 303.8 120.1 115.9 64.8 15.5 98.35 99.31 91.61 95.72 97.73 92.15 99.93 95.76 91.90 92.01 82.35 91.61

(%) 99.96 80.81 95.12 98.88 96.55 99.37 95.83 97.18 78.88 99.78 96.33 98.62 81.57

1.94 8.41 1.45 2.64 1.13 1.47 16.4 1.73 1.23 12.13 1.07 4.51

(mg/L) 4.03 0.11 1.94 3.28 3.30 1.91 0.35 2.05 0.78 3.82 2.01 1.77 0.23

As(III) As(III) Fetot

0.06 0.14 0.08 0.23 0.23 0.04 0.46 0.08 0.24 0.13 0.98 0.11

(mg/L) 0.04 0.32 0.09 0.03 0.07 0.02 1.55 0.01 0.07 0.08 0.10 0.12 0.18

Mn

80

60

Mg

60

SO

4

+C l

+ Ca

80

40

40

20

20 SO4

3

O HC

K

60

80

+

+ Na

80

CO

3

Mg

60

40

40 20

20 80

60

40

20

Ca

20

40

60

80

Na HCO3

Cl

Figure 3. Piper diagram showing the dominance of Ca-Mg-HCO3 water type in TAP groundwaters in Nawalparasi. 5 y = 0.1408x - 0.8508

y = 0.0069x + 0.9303

4

R2 = 0.5883

Fetot (mg/L)

Fetot (mg/L)

5

3 2

4

R2 = 0.5638

3 2 1

1 a

b 0

0 0

100

200

300

10

400

15

Astot (µg/L)

20

25

30

35

DOC (mg/L)

Figure 4. Plots showing the relationship between: (a) Astot, and Fetot and (b) DOC and Fetot in TAP groundwaters in Nawalparasi.

Astot (µg/L)

250

30

a

y = 0.578x - 158.95 R2 = 0.5393

DOC (mg/L)

300

200 150 100 50 0 200

300

400

500

y = 0.0281x + 5.4257 R2 = 0.6763

20 15 10 200

600

b

25

300

400

500

600

HCO3 (mg/L)

HCO3 (mg/L)

⫺ Figure 5. Plots showing the relationship between: (a) HCO⫺ 3 and Astot; and (b) HCO3 and DOC in TAP groundwaters in Nawalparasi.

45

3,0

y = 0.0056x + 0.2476

2,5

R2 = 0.8946

DOC (mg/L)

NH4+ (mg/L)

3,5

2,0 1,5 1,0

35

y = 0.0346x + 15.397

30

R2 = 0.7301

25 20 15

0,5

a

b 10

0,0 0

100

200

300

400

500

0

100

As(III) (µg/L)

200

300

400

500

As(III) (µg/L)

Figure 6. Bivariate plots showing the relationship of As(III) with NH⫹ 4 and DOC in TAP groundwaters in Nawalparasi. Fetot, Mn (mg/L)

Astot (µg/L) 100

200

300

400

0

500

10

10

20 30 40 50 60

Depth (m)

0

WHO Safe Drinking Water Limit

0 Interim Nepalese Drinking Water Standard

Depth (m)

0

Mn

10

15

20 Fetot

20 30 40 50

a

5

b

60

Figure 7. Variation with depth the concentration of (a) Astotal and (b)Fetotal and Mn in TAP groundwaters in Nawalparasi.

presence of Fe and Mn at elevated concentration, together with the predominance of As(III) in the groundwater. Elevated HCO⫺ 3 concentrations result primarily due to the oxidation of organic matter (Mukherjee & Bhattacharya 2001, Bhattacharya et al. 2002), while low SO42⫺ concentrations result due to reduction of sulfate. Strong correlation between DOC and HCO3 indicate the abundance of degradable organic matter (Bhattacharya 2002, Bhattacharya et al. 2004). The presence of high DOC levels coupled with dominance of As(III) in groundwater suggest strong anoxic conditions caused by microbially mediated reduction of organic matter. These co-relations strongly support the hypothesis of reductive dissolution of Fe-Oxyhydroxides as the main mechanism of mobilization of As in groundwater.

6

MITIGATION OPTIONS

To co-ordinate and streamline all the activities related to As of different agencies under single umbrella ‘The National Steering Committee on Arsenic (NSCA)’ has been formed with 20 members representing different governmental, non-governmental and donor agencies working in the field of water, sanitation and health sector. Information on As should be disseminated properly to avoid imminent danger. Therefore to provide uniform flow of information, IEC and training materials that are suitable in the context of Nepal have already been printed and are being distributed in the hot spot areas. Since training is an effective way to disseminate information on As, a network of trainers has already been established by imparting training to more than 300 staffs of DWSS 46

and about 140 members from other organizations engaged with arsenic mitigation activities. These trainees, especially the frontline workers will go to the affected areas to create awareness on As problem and deal with mitigation options. Blanket testing by As-field test kit is the easiest way for screening to find out As free sources nearby for tubewell switching. Hence, DWSS in collaboration with UNICEF and WHO is conducting blanket testing program in 10 arsenic affected districts of TAP. In the absence of As free source nearby, the only available option is the treatment of water either at the point of entry or at the point of use to meet the drinking water standard. After getting the complete blanket test result, As treatment methods, which are simple, effective, affordable and socially acceptable treatment options will be provided to the affected communities in hotspot areas. A few institutions have already studied the simple options namely 3-gagri filter, arsenic biosand filter etc. These filters use locally available materials. However, such treatment options should be used for short-term remediation only. In long term plan the affected people should be provided with As free water. In Bangladesh, the study conducted by JICA/AAN shows that 23 out of 51 dugwells and 38 out of 243 deep tubewells were found to have arsenic concentration exceeding the limit of 50 ␮g/L (JICA/AAN 2004a, b). It shows that deep tubewell and dugwell waters are not necessarily always safe hence these wells should compulsorily be tested before recommending them as a safe source. It may be equally applicable in the context of Nepal also.

7

CONCLUSIONS

The detection of As in groundwater of TAP in southern Nepal has raised concern about health risk for about one fourth million people. Positive correlation between DOC, HCO⫺3 , FeTotal and AsTotal in groundwater indicate that As is mobilized primarily due to the reductive dissolution of Feoxyhydroxide in the presence of organic matter in the sediments of TAP. Blanket As testing by field kit is the easiest way to find out As free source nearby for tubewell switching. In the absence of As free source, the only available option is the treatment of water either at the point of entry or at the point of use to meet the drinking water standard. Treatment methods namely 3-gagri filter and arsenic biosand filters can be installed in the hotspot areas as short-term remedial options. However, in the long-term plan affected communities should be provided with As free water by tapping sources from springs, rain water harvesting, treatment of water from rivers etc.

ACKNOWLEDGEMENTS This study was carried out as a part of M.Sc. thesis with financial support by KTH. We acknowledge DWSS, His Majesty’s Government of Nepal for providing all the logistic support during the field work. We would like to thank Ann Fylkner, Monica Löwen (at the laboratories of Land and Water Resources Engineering, Royal Institute of Technology) and Joyanto Routh, Thomas Hjorth (Stockholm University) for their help in doing chemical analysis. We would also thank D. Chandrashekharam for his constructive comments on an earlier draft of this manuscript.

REFERENCES Bhattacharya, P. 2002. Arsenic contaminated groundwater from the sedimentary aquifers of South-East Asia. Groundwater and Human Development, Proc. XXXII IAH and VI ALHSUD Congress, Mar del Plata, Argentina, E. Bocanegra, D. Martinez and H. Massone, 357–363. Bhattacharya, P., Jacks, G., Ahmed, K.M., Khan, A.A. & Routh, J. 2002. Arsenic in groundwater of the Bengal Delta Plain aquifers in Bangladesh. Bull. Env. Cont. Toxicol. 69: 538–545. Bhattacharya, P., Tandukar, N., Neku, A., Valero, A.A., Mukherjee, A.B. & Jacks, G. 2003. Geogenic arsenic in groundwaters from Terai alluvial plain of Nepal. Jour. de Physique IV France 107: 173–176.

47

Bhattacharya, P., Ahmed, K.M., Broms, S., Fogelström, J., Jacks, G., Sracek, O., von Brömssen, M. & Routh, J. 2004. Mobility of arsenic in groundwater in a part of Brahmanbaria district, NE Bangladesh. Managing Arsenic in the Environment: From Soil to Human Health, R. Naidu, E. Smith, L. Smith, J. Smith, and P. Bhattacharya (eds), CSIRO Publishing, Melbourne, (in press). GWRDB-UNDP 1989. Shallow groundwater exploration in the Terai. Nawalparasi District (West). Technical Report No. 5, Kathmandu, Nepal, 21p. JICA/AAN 2004a. Japan International Cooperation Agency/Asia Arsenic Network Arsenic Mitigation Project. Arsenic contamination of Deep Tubewells in Sharsha Upazila, Bangladesh, Report 1. JICA/AAN 2004b. Japan International Cooperation Agency/Asia Arsenic Network Arsenic Mitigation Project. Water quality and Follow-up survey on Arsenic contamination of Dugwells in Sharsha Upazila, Bangladesh, Report 2. Meng, X. & Wang, W. 1998. Speciation of arsenic by disposable cartridges. 3rd Int. Conference on Arsenic Exposure and Health Effects, San Diego, CA. Meng, X., Korfiatis, G.P., Christodoulatos, C. & Bang, S. 2001. Treatment of arsenic in Bangladesh well water using a household co-precipitation and filtration system. Water Resources. 35: 2805–2810. Mukherjee, A.B. & Bhattacharya, P. 2001. Arsenic in groundwater in the Bengal Delta Plain: Slow Poisoning in Bangladesh. Env. Rev. 9: 189–220. NSCA 2001. Nepal’s Interim Arsenic Policy Preparation Report. Draft report, pp 29. Tandukar, N. 2000. Arsenic Contamination in Groundwater in Rautahat District of Nepal – An Assessment and Treatment, Unpublished M.Sc. Thesis, Institute of Engineering, Lalitpur, Nepal. Tandukar, N., Bhattacharya, P. & Mukherjee, A.B. 2001. Preliminary assessment of arsenic contamination in groundwater in Nepal. Book of Abstracts, Arsenic in the Asia-Pacific Region Workshop, CSIRO, Adelaide, Australia, 103–105. Valero, A.A. 2002. Arsenic in groundwater of alluvial aquifers in Nawalparasi and Kathmandu districts of Nepal: Extent of contamination, genesis and aspects of remediation. TRITA-LWR Master Thesis, 02-12, ISSN 1651-064X, KTH, Stockholm, Sweden, 57p.

48

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

High arsenic concentrations in mining waters at Kan
k, Czech Republic A. Kop
riva, J. Zeman & O. Sracek Institute of Geological Sciences, Faculty of Science, Masaryk University, Brno, Czech Republic

ABSTRACT: Ore deposit Kan
k is located southeast of Prague, Czech Republic, close to historical mining town Kutná Hora (Kutennsberg) where the mining of silver ores started in 14th century. In 1991, the mine at Ka
nk was closed. In November 2001, catastrophic event occurred after mine water reached Dedicna stola drift when the concentrations of iron suddenly increased to 7000 mg/L, and that of arsenic increased to 60 mg/L. Based on the study of mine water geochemistry, the conceptual model of water chemistry evolution can be summarized as follows: (a) During mining activities and de-watering of mine, oxidizing conditions prevailed. Arsenopyrite was oxidized and released. (b) During flooding arsenic was released from dissolved minerals, concentrations of both arsenic and iron became higher, and stratification evolved. (c) When water level reached de-watering drift, lower concentrated layer is discharged for relatively short period. (d) This period is followed by outflow of highly mineralized water with extremely high concentrations of arsenic and other species.

1

INTRODUCTION

Arsenic is one of the most toxic elements found in nature. Many sources of arsenic are geogenic (Smedley & Kinniburgh, 2001; Ahmed et al. 2004). In that case arsenic enrichment in environment occurs as a consequence of oxidation of sulfide minerals, reductive dissolution of ferric oxide and hydroxides with adsorbed arsenic, hydrothermal activities, and evaporation enrichment of arsenic dissolved in water (Welch et al. 2000). Recently, the safe limit for arsenic concentration in drinking water was set by WHO to 0.01 mg/L. This was a consequence of arsenic poisoning in the Bengal Delta Plain (BDP) in Bangladesh and West Bengal in India. However, in some cases arsenic enrichment in environment can be caused by human activities like mining. In that case, primary arsenic sulfide minerals like arsenopyrite and As-rich pyrite are oxidized under oxidizing conditions during mining and released arsenic is stored in secondary minerals like scorodite, farmacosiderite, bukovskyite etc. or adsorbed on ferric oxide and hydroxides. After a mine closure and flooding of mining works, arsenic is released into ground water as a consequence of redox conditions changes. Kan
k is located southeast from Prague, Czech Republic, in mining district around historical mining town Kutna Hora (Kutennberg), which is an example of such evolution. Here mining activities were terminated in 1991 and flooding of the mine started. Concentrations of arsenic in some cases reached ⬎100 mg/L and this makes the site unique even at world scale. The purpose of this text is to discuss conditions leading to such extreme concentrations. 2

GEOLOGICAL CONDITIONS

Ore deposit Kan
k opened by the mine Kutna Hora is located close to the historical mining town Kutna Hora (Kutennsberg), Fig. 1. Here mining activities started in 14th century and silver was principal target of mining. After depletion of silver, mining activities were focused on Pb-Zn mineralization. 49

Figure 1.

Geographic location of Kutna Hora, Czech Republic.

The ore deposit is located in the north of Kutna Hora mining district. The basement is formed by Proterozoic and Lower Paleozoic rocks overlain by sediments of Bohemian Cretaceous Formation and Quaternary sediments. Host rocks of hydrothermal mineralization of Variscian age (320–320 Ma) are metamorphic rocks of Kutna Hora Formation, predominately formed by biotitic gneisses and migmatites. On the top of metamorphic complex at northeast of Kutna Hora town there are blocks of calcareous sandstones and organodetritic limestones covered by Quaternary loess and soils. The mineralization has a polymetalic character and is represented by carbonate and quartz veins with sulfides of Fe, As, Pb, Zn, Cu, Sb, Sn, and Ag. Most important As-bearing mineral is arsenopyrite, less significant are As-pyrite and pyrrhotite. Local arsenopyrites exhibit significant variations in As content, varying from 31.6 to 34.2wt% with minor Bi content (about 0.14 wt%). There is an assemblage of secondary As-minerals present in waste rock piles and in abandoned underground mining corridors (bukovskyite, kankite, zykaite, skorodite etc.). First indications of mining activities at Kutna Hora are from 14th century. Mining was stopped in 1904, but re-started during the World War II. After the war mining continued only in the north part of the ore deposit. In 1956, Turkanska shaft reached a depth of 550 m bgl and 7th floor was opened. Mining activities ended completely in 1991 and controlled flooding of mines has started. From 1993 regular monitoring of mining water quality has been performed. In early September 2001 water level in mining shaft reached Dedicna stola drift (at 210 m a.s.l.). In November 2001 there was a unusual increase in the concentrations of dissolved component (Fe, SO4, Zn etc.) and two small surface streams in the proximity of the mine were contaminated. This resulted in pumping of water from shaft and an emergency treatment plant was built. Currently water level in the shaft is maintained below Dedicna stola drift level and water is pumped from deeper zone of the shaft. However, it is only a temporary solution and several problems, including long-term deposition of treatment slurry are still unresolved.

3

EVOLUTION OF WATER CHEMISTRY DURING FLOODING

Mining activities severely disturbed the natural ground water flow conditions. Large cone of depression was formed around mine as described by Younger et al. (2002). During mining period 50

250 200

210.3 masl – decant to outflow pathway

Mine Water Level (m n. m.)

150 100 50 0 -50 -100 -150 -200 -250 91

Figure 2.

92

93

94

95

96

97 Year

98

99

00

01

02

Rise of water level in shaft collar at Kan
k from 1992 to 2002.

the following inflows were measured: 1st floor ⫺20 L/m, 2nd floor ⫺16 L/m, 3rd floor ⫺205 L/m, 4th floor ⫺80 L/m, 5th floor ⫺90 L/m, 6th floor ⫺9 L/m. Total inflow to the mine during mining activities was 420 L/m. During flooding there was a gradual increase of water level in the shaft (Fig. 2). The rate of water level recovery was initially linear, but decreased at later stage. Evolution of pH, and dissolved concentrations of Fe and Mn during flooding in water from mining shaft are shown in Figure 3. Evolution of dissolved concentrations of Zn, As, and SO4 are in Figure 4. During the early stages of flooding (1st stage) until 1993), low concentrations (less than 5 mg/L TDS) dominated. The chemistry of water in the shaft was influenced by recharge of relatively fresh ground water from fractures in the walls of the shaft. In the 2nd stage from 1994, concentration of TDS increased to values about 15 g/L and concentrations of individual dissolved components were correspondingly higher. This relatively fast increase in concentrations of TDS was followed by gradual decrease during this period of mine flooding with marked seasonal variations. 3rd stage started after Dedicna stola drift was reached by rising water table. For more than two month there was almost no visible change in mine water geochemistry. 4th and final stage started after this period and is characterized by dramatically increased concentrations of dissolved components. This stage of mine water geochemical evolution lasts until now. Concentrations of dissolved iron and arsenic sampled close to water table were relatively low and stable at the beginning of flooding. Later there was gradual increase of total iron concentration to 2700 mg/L, followed by decrease and relative stabilization at about 1500 mg/L at the end of 1990s. Behavior of dissolved arsenic was generally related to the behavior of iron, but there were more significant variations. Concentrations fluctuated between 15 and 60 mg/L, but reached ⬎100 mg/L in several sampling events. At the same period, pH values increased from initial values ⬍2.0 to values ⬎3.0. In November 2001, disastrous event occurred because concentrations of Fe suddenly reached 7000 mg/L, and As concentrations stabilized around 60 mg/L. Several surface streams close to the mine were at high risk of contamination. Emergency plant for arsenic treatment was immediately built. 4

CONCEPTUAL MODEL OF GEOCHEMICAL EVOLUTION

Conceptual model of water chemistry evolution at Kan
k can be summarized as follows (Fig. 5): (a) During the mining activities and de-watering of mine, oxidizing conditions prevailed. Arsenopyrite was oxidized in reaction like 51

5 4

3

3 pH

4

2

2

1

1

0 93 94 95 96 97 98 99 00 01 02 03 Year

0 01

8000

8000

7000

7000

6000

6000

5000 4000 3000

5000 4000 3000

2000

2000

1000

1000 0 01

0 93 94 95 96 97 98 99 00 01 02 03 Year

250 200 Mn [mg/L]

200 Mn [mg/L]

02 Year

250

150 100

150 100 50

50 0 93 94 95 96 97 98 99 00 01 02 03 Year

Figure 3.

02 Year

Fe [mg/L]

Fe [mg/L]

pH

5

0 01

02 Year

Evolution of pH, Fe, and Mn concentrations during flooding of the mine.

Then ferric iron under oxidizing conditions precipitated in minerals such as amorphous ferric hydroxide, Fe(OH)3(am). Released arsenic was adsorbed on ferric oxides and hydroxides. A part of dissolved arsenic precipitated in the form of secondary minerals like scorodite, bukovskyite, and kankite. This means that arsenic and iron were relatively immobile under oxidizing conditions, which prevailed during the mining activities. Behavior of sulfate was more conservative then behavior of arsenic and iron, but also a part of sulfate was probably stored in minerals like gypsum, CaSO4⭈2H2O, and jarosite, KFe3(SO4)2(OH)6. (b) After closure, mine is flooded and conditions at the bottom of mining shaft become increasingly reducing. Arsenic is released and/or desorbed from dissolving secondary minerals and concentrations of both arsenic and iron become increasingly high. During this period, water chemistry 52

Zn [mg/L]

160

70

140

60

120

50 40 30

100 80 60

20

40

10

20 Year

0 93 94 95 96 97 98 99 00 01 02 03 Year

02

25 000 20 000 SO42– [mg/l]

20 000 15 000 10 000

15 000 10 000 5 000

5 000 0 93

02 Year

25 000

SO42– [mg/l]

1000 800 600

80

0 01

Figure 4.

1800 1600 1400 1200

400 200 0 01

As [mg/L]

Zn [mg/L]

As [mg/L]

1800 1600 1400 1200 1000 800 600 400 200 0 93 94 95 96 97 98 99 00 01 02 03 Year

95

97 99 Year

01

03

0 01

02 Year

Evolution of Zn, As, and SO42⫺ concentrations during flooding of the mine.

stratification in the mine shaft develops – high concentrations layer at the bottom of shaft and low concentrations layer close to water level due to the processes like recharge of ground water with low mineralization flowing from fractures intersecting walls of the mine shaft above water level. Sampling of water close to water table yields samples with relatively low concentrations of Fe, sulfate and arsenic. However, when deep sampling is performed, concentrations of dissolved constituents are much higher. Thus, sampling of shallow water in the shaft gives a false impression of low concentrations and underestimates potential danger for contamination in future. (c) When water level in the shaft reaches de-watering drift, the layer of mine water with lower concentrations of dissolved species is discharged and therefore there is no apparent change in water geochemistry. (d) When the interface between relatively fresh water and high mineralization water reaches outflow drift, there is almost sudden change in water quality. This results in the flow of highly mineralized water with extremely high concentrations of iron and arsenic from deeper zone to the de-watering drift. 53

Figure 5. Conceptual model of mine water geochemical evolution during flooding of mine at Kan
k (see text for more explanation).

This behavior was also observed in coal mines (Nutall & Younger, 2004). Furthermore, there are data from other flooded mines in Czech Republic – for example, in Zbysov in Southern Moravia, uranium mine at Dolni Rozinka and Brzkov, and polymetallic and uranium mine at Pr
íbram (J. Zeman & A. Kopr
iva, unpublished data) where discrete-depth sampling has been performed. Stratification with high concentrated layer of mine water under more reducing conditions in depth has been generally found.

5

CONCLUSIONS

There are extremely high concentrations of dissolved As concentrations (⬎60 mg/L) in mining waters at Kank mining district close to Kutna Hora in Czech Republic. After closure of mine in 1991, there was gradual flooding of mining shaft. Concentrations of dissolved constituents sampled close to water table were relatively low initially. When water level in the shaft reached Dedicna stola de-watering drift, there was a sharp increase of dissolved concentrations after relatively short 54

initial time lag period. A conceptual model of geochemical evolution has been developed: during mining oxidizing conditions prevail and As is adsorbed on ferric oxide and hydroxides or incorporated into secondary minerals; at early stage of flooding, there is release of arsenic from dissolving secondary minerals and stratification in mining shaft with increasing concentrations with depth develops; and in late stage of flooding water level reaches de-watering drift and dissolved concentrations become extremely high. These findings are applicable for flooding of any mines and future work will be focused on better understanding of the development of stratification.

REFERENCES Ahmed K.M., Bhattacharya P., Hasan M.A., Akhter S.H., Alam S.M.M., Bhuyian M.A., Imam M.B., Khan A.A. & Sracek O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh: An overview. Appl. Geoch. 19: 181–200. Nutall C.A. & Younger P.L. 2004. Hydrogeochemical stratification in flooded underground mines: an overlooked pitfall. J. Contam. Hydrology 69: 101–114. Smedley P. & Kinniburgh D.G. 2002. A review of the source, behavior and distribution of arsenic in natural waters. Appl. Geochem. 17: 517–568. Welch A.H., Westjohn D.B., Helsel D.R. & Wanty R.B. 2000. Arsenic in Groundwater of the United States: Occurrence and Geochemistry. Groundwater 38(4): 589–604. Younger P.L., Banwart S.A. & Hedin R.S. 2002. Mine Water, Hydrology, Pollution, Remediation, Kluwer Academic Publishers.

55

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Natural arsenic in the groundwater of the alluvial aquifers of Santiago del Estero Province, Argentina Prosun Bhattacharya, Mattias Claesson & Jens Fagerberg Groundwater Arsenic Research Group, Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden

Jochen Bundschuh, Angel del R. Storniolo, Raul A. Martin & Juan Martin Thir Facultad de Ciencias Exactas y Tecnologias, Universidad Nacional de Santiago del Estero (UNSE), Santiago del Estero, Argentina

Ondra Sracek Institute of Geological Sciences, Faculty of Science, Masaryk University, Brno, Czech Republic

ABSTRACT: Natural occurrences of arsenic has been documented in groundwater of the shallow aquifers of the Chaco-Pampean Plain, Argentina. The distribution of arsenic and mechanisms of its mobilization in the shallow alluvial aquifers was investigated around the city of Santiago del Estero in Northwestern Argentina in order to provide an insight into the complex hydrological and geochemical conditions that yields high As concentrations in groundwater. Significant spatial variations of total arsenic (Astot) concentrations were observed with an average value of 743 ␮g/L. Arsenate was a dominant species in most samples. Average concentrations of Al, Mn, and Fe were 360 ␮g/L, 574 ␮g/L, and 459 ␮g/L, respectively. The 7M HNO3 extraction of sediments and volcanic ash-layer indicated AsNO3 concentrations ranging between 2.5–7.1 mg/kg. AsNO3 indicated a significant positive correlation with MnNO3, AlNO3, and FeNO3. Oxalate extractions revealed significant fractions of As (Asox) in the sediments (0.4–1.4 mg/kg) and a dominance of oxalate extractable Al- and Mn. Speciation calculations indicate that Al oxide and hydroxides have the potential to precipitate in the groundwater, suggesting that As adsorption processes may be to some extent controlled by Al oxides and hydroxides. Mobility of As at local scale seems to depend on high pH values, related to the dissolution of carbonates driven by cation exchange, and dissolution of silicates. There is a clear relationship of As with F, V, B and Si, suggesting their common origin in volcanic ash layer. Preliminary conceptual model of arsenic input includes release of As and Al from dissolution of volcanic ash layer, precipitation of Al oxides and hydroxides followed by adsorption of As on Al and Fe phases in sediments, and release of As under high pH conditions.

1

INTRODUCTION

Arsenic (As) is a natural inorganic contaminant in drinking water, which is known to have caused serious environmental health problems globally (Bhattacharya et al. 2002a, Smedley & Kinniburgh 2002). Elevated concentrations of As from geogenic sources are reported in groundwaters in different parts of the world such as Argentina, Bangladesh, China, Nepal, Mexico, Vietnam, and United States among others (Bhattacharya et al. 2002a, Smedley & Kinniburgh 2002, Bhattacharya et al. 2004). Natural occurrences of As has been documented in groundwaters in Argentina from the alluvial aquifers of the Chaco-Pampean Plain, where a population of approximately 1.2 million mostly in rural settlements are exposed to As from local drinking water sources. The concentrations of As in the groundwater are mostly above the limit of safe drinking water (10 ␮g/L; WHO 57

2001) as well as the local drinking water standard of 50 ␮g/L. The first symptoms of arsenic related diseases were detected in 1983 within the counties of La Banda and Robles, and in the following year investigations by the government agencies confirmed elevated arsenic concentrations (above 1000 ␮g/L) in groundwater of the shallow aquifers around the provincial capital of Santiago del Estero in North-western Argentina (Martin 1999). In the recent years, chemistry of groundwater of the shallow aquifers within selected areas of the alluvial cone of the river Río Dulce of Santiago del Estero province (Bejarano & Nordberg 2003, Claesson & Fagerberg 2003, Bundschuh et al. 2004) have been studied in order to understand the mechanisms of arsenic mobilization in these aquifers. The aim of this study was to investigate the distribution of As in the groundwater of shallow aquifers located within the alluvial cone of the river Río Dulce around the city of Santiago del Estero and the associated shallow sediments in order to develop a better understanding of the complex hydrogeological and geochemical conditions that are responsible for the mobilization of As in groundwater.

2 2.1

LOCATION AND GEOLOGICAL SETTING The study area

Santiago del Estero is the provincial capital of the province with the same name. It is located on the dry, western part of the Chaco plain in northwestern Argentina at an altitude of about 200 meters above sea level (Fig. 1). The study area is located to the east of the capital on the alluvial deposits formed by the Río Dulce (hereinafter referred as the Río Dulce alluvial cone), covering an area of approximately 2000 km2. The area is rural, but densely populated in comparison with surrounding countryside due to its fertile soils and irrigation systems built up by channels from the Río Dulce. Small agricultural settlements dependent on artificial irrigation are common throughout the Río Dulce cone. Hot, rainy summers lasting from November to March and very dry winters from April to October characterize the climate. The average annual precipitation (1938–90) is 532 mm. Evapotranspiration is very high, especially in summer period. The area was originally forested but due to intensive timber harvesting only limited forest areas remain. Vegetation generally consists of low bushes and ground vegetation. The terrain is very flat with few shallow depressions resulting from historical flow-paths of the river. Strong winds are common and carry a lot of dust during winter when binding surface water is scarce. 2.2

Geological and hydrogeological characteristics

Río Dulce alluvial cone is limited to the west by the Huyamampa fault. A sequence of 30 m Pleistocene Pampean loess is found to the west of this fault comprising dispersed volcanic ash and calcareous crusts, which is underlain by lower Pliocene and green Miocene clays with a thickness of 70 m. Holocene to recent sediments of the Río Dulce cone (Post-Pampean formation) occur to the east of the fault, and the thickness of the fluvial and aeolian sediments is nearly 150 m close to the fault margin and pinches out about 50 km towards the east. A sequence of alternating layers of gravel, sand, silt and clay is deposited in discordance with the underlying Pliocene sediments. The coarser sediments represent the deposits of the palaeo-channels of the river Río Dulce, which form a multi-layered aquifer system in the region. Santiago del Estero is located just south of where Río Dulce passes the fault of Huyamampa (Fig. 2). The fault runs from north to south and the river passes from northwest to southeast. West of the fault Pliocene and Miocene clays are covered by approximately 30 m of Pampean loess (Bundschuh et al. 2004). Southeast of the fault, the river have deposited alluvial sediments making up the Río Dulce cone. Depth to ground water table ranges from 1 to 6 m. The upper-most aquifer, the one of major interest in this study, reaches an approximate depth of 15 m (Martin 1999). An 58

Figure 1. Digital elevation model of the southern part of South America with location of the project area “Río Dulce alluvial cone” located near the city of Santiago del Estero in North western Argentina (digital elevation model modified from PIA03388 image; http://photojournal.jpl.nasa.gov). Other areas with groundwater arsenic are also shown.

important fraction of groundwater recharge to all aquifers, including the shallow one, takes place in a limited section of coarse material along the Huyamampa fault in the northwestern part of the study area. Here all aquifers in the alluvial cone are connected to the surface and are recharged by infiltrating river and surface water. The aquifers are considered to be semi confined with little interactions among them. Major groundwater flow is horizontal in the separate aquifers. The upper-most aquifer consists of aeolian and fluvial sediments (Bundschuh et al. 2004) and has important recharge from the river and surface not only near the Huyamampa fault, but all over its area. Irrigation channels are non-lined which may lead to significant infiltration losses. Local groundwater flow patterns yielding long residence times in natural depressions are likely to be of importance for groundwater chemistry. General direction of flow in the upper aquifer is towards SE (Fig. 2) and several localized discharge areas occur in topographic lows. However, data on local groundwater flow patterns and hydraulic conductivity of the upper-most aquifer are very limited. 59

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Figure 2. Detailed map of the Río Dulce alluvial cone, the location of the sampled wells and the generalized patter of groundwater flow within the alluvial cone.

3 3.1

MATERIALS AND METHODS Groundwater and sediment sampling

Groundwater samples were collected during September–October 2002 from the counties Robles and La Banda (Fig. 2), both within a distance of not more than 50 kilometers from the city of Santiago del Estero. Forty well sites were selected where groundwater was mainly abstracted by hand-pumped tube-wells penetrating the shallow aquifers, that reached a maximum depth of 12 m. Sampling points are shown in Figure 2. The well positions at each of the sampling site were determined using Global Positioning System (GPS). The values of pH, redox potential (Eh), temperature and electrical conductivity of groundwater were measured in the field. Water samples collected from each well involved: (i) filtered samples for alkalinity and major anion analysis; (ii) filtered, acidified samples for major cation and trace elements analysis; (iii), sample for DOC analysis, and (iv) filtered through a Disposable Cartridge® for field separation of As(V) and As(III). Sediment samples were collected from two sites, Balsamo Cuatro Horcones (CH 47) and Nuevo Libano (NL 30) (Fig. 2) up to a depth of 1.2 m using an auger drill in the immediate vicinity of tube wells to study the sediment-groundwater interactions. Groundwater and the sediment were analysed following the procedure outlined by Bhattacharya et al. 2001, 2002b. 3.2

Analytical methods

⫺ Anions such as Cl⫺ and SO2⫺ 4 were analyzed by a Dionex 120 ion chromatograph, and NO3 and 3⫺ PO4 were analyzed using Tecator AQUATEC 5400 analyzer at wavelengths 540 nm and 690 nm, respectively. The major and trace metals were analyzed on a Perkin Elmer Elan 6000 ICP-MS. As(V) was calculated as a difference between total As and As(III) in the samples. Certified standards, SLRS-4 (National Research Council, Canada) and GRUMO 3A (VKI, Denmark) and synthetic

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Figure 3.

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Representation of major ion chemistry in groundwater samples plotted on a Piper diagram.

chemical standards prepared in the laboratory, and duplicates were analyzed after every 10 samples during the runs. Trace element concentrations in standards were within 90–110% of their true values. In case of wider variations, the standards were recalibrated and the preceding batch of 10 samples reanalyzed. Relative percent difference between the original and duplicate samples were within ⫾10%. Dissolved organic carbon (DOC) in the water samples were determined on a Shimadzu 5000 TOC analyzer (0.5 mg/L detection limit with a precision of ⫾10% at the detection limit).

4 4.1

RESULTS Groundwater chemistry

Groundwater pH ranged between 6.4 and 9.3 with an average of 7.6. Field measured redox potential ranged from ⫺60 to ⫹348 mV with an average value of ⫹153 mV. Electric conductivity (EC) ranged between 804 and 9800 µS/cm with an average is 2422 ␮S/cm. Major ion composition indicated Na⫹ (average concentration 427 mg/L) and HCO⫺ 3 (581 mg/L) as the dominating ions in groundwaters (Fig. 3). DOC concentrations in groundwater varied between below detection limit and 18.2 mg/L with an average of 7.6 mg/L. Total arsenic (Astot) concentration indicated considerable spatial variations with an average of 743 ␮g/L. Some wells exhibited extremely high values, reaching a maximum of 14,969 ␮g As/L (Fig. 4b). Speciation of As indicated the dominance of As(V) with average concentration of 617 ␮g/L, while the concentration of As(III) averaged around 125 ␮g/L. Among the other trace elements, dissolved Al concentrations were low (average 360 ␮g/L, median 17 ␮g/L), while the concentrations of Mn (average 574 ␮g/L, median 128 ␮g/L) and Fe (average 459 ␮g/L, median 140 ␮g/L) were higher. These groundwaters were characterized by high Si concentrations (average 28.1 mg/L). Fluoride concentrations were elevated (average 2.55 mg/L, median 1.26 mg/L), which also exceeded the WHO limit for safe drinking water (1.5 mg/L; WHO 1993) in 16 out of the 40 sampled wells. 61

8

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Figure 4. Bivariate plots showing the correlation of AsNO3 with: a) FeNO3 (䊐 – dashed line; R ⫽ 0.79) and Al (䉱 – solid line; R ⫽ 0.79), and b) Mn NO3 (R ⫽ 0.76) in the shallow aquifer sediments.

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Figure 5. Sequential leaching of As from the sediment samples from: left: Balsamo Cuatro Horcones (CH 47); right: Nuevo Libano (NL 30). Sediments were sequentially leached by de-ionized water (DIW), bicarbonate (HCO3), acetate and oxalate.

4.2

Sediment chemistry

Geochemical investigations have revealed considerable enrichment of arsenic in shallow aquifer sediments. Extraction of the sediments samples collected at two sites (NL30 and CH47, Fig. 2) by 7M HNO3 revealed AsNO3 concentrations in the range between 2.5–7.1 mg/kg which is higher than the average As concentrations in soils and sediments (Taylor & McLennan 1985). The volcanic ash-layer also had appreciable AsNO3 content (3.0 mg/kg). A significant positive correlation (Fig. 4) was observed between AsNO3 and the concentrations of MnNO3, Al NO3, and FeNO3 (R ⫽ 0.76–0.79). Sequential leaching of sediment samples was performed using deionized water (DIW), bicarbonate (HCO3), acetate and oxalate media, which extracted As at varying concentrations (Fig. 5). Oxalate extractions reveal significant fractions of extractable As (Asox) in all the samples ranging between 0.4–1.4 mg/kg (Fig. 5). Comparison of Fe, Al and Mn concentrations extracted by oxalate clearly indicated dominance of Al- and Mn-oxides and -hydroxides as compared to Fe-oxides and hydroxides (Fig. 6). Much higher concentrations at site CH47 seem to be related to the preferential dissolution of volcanic ash layer located at this site at about 1 m depth. 4.3 Speciation calculations Results of saturation indices (SI) calculations for selected minerals calculated with program MINTEQA2 are listed in Table 1. They show that crystalline forms of Al oxide and hydroxides such as gibbsite are stable in groundwater, implying that As adsorption processes may be to some 62

Oxalate leached Fe, Al, Mn (mg/kg) 100 200 300 400 500 600

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Figure 6. Trends in the distribution of oxalate extractable Fe, Al and Mn with depth of the sediment samples from: left: Balsamo Cuatro Horcones (CH 47); right: Nuevo Libano (NL 30).

Table 1. Results of speciation modeling using the MINTEQA2 program (Alison et al. 1991) and the SI values for selected Al, Fe, and Mn phases for ground water samples from five wells.

Mineral

NL30 (Astot ⫽ 8083 ␮g/L)

SC8 (Astot ⫽ 1574 ␮g/L)

23 (Astot ⫽ 669 ␮g/L)

24 (Astot ⫽ 36 ␮g/L)

16 (Astot ⫽ 23 ␮g/L)

Al(OH)3(a) Gibbsite Fluorite Fe(OH)3(a) Goethite MnOOH Rhodochrosite Siderite SiO2 (a) Vivianite

SI ⫺1.98 0.74 1.09 0.31 6.18 ⫺8.21 ⫺0.32 0.57 ⫺0.57 ⫺1.00

SI ⫺2.68 0.03 ⫺0.33 2.08 7.95 ⫺4.19 ⫺1.14 ⫺2.53 ⫺0.67 ⫺10.37

SI ⫺2.50 0.27 0.81 2.18 8.0 ⫺1.38 ⫺0.21 ⫺4.18 ⫺0.47 ⫺21.04

SI ⫺3.04 ⫺0.33 ⫺1.46 1.87 7.73 ⫺2.94 0.28 ⫺2.54 ⫺0.60 ⫺13.65

SI ⫺1.35 1.37 n.a. ⫺1.58 4.29 ⫺9.36 ⫺1.19 ⫺0.81 ⫺0.60 n.a.

n.a. – not available; (a) – amorphous phase.

extent controlled by Al mineral phases. This is consistent with an important role of Al mineral phases in oxalate extractable fraction. However, also ferric minerals like goethite are stable and may be important As adsorbents.

5

DISCUSSION AND CONCLUSIONS

Potential primary source of both As and Al is volcanic ash layer, which comprises very soluble volcanic glass. Principal Component Analysis (Bhattacharya et al. 2004, submitted) shows the relation of As with F, V, B and Si; most likely due to their common origin in volcanic ash, indicating its importance as a source of As in shallow groundwater. This is exhibited by a fairly high correlation (Fig. 7a–d) observed between the concentrations of Astot with F⫺ (R2 ⫽ 0.43; p ⬍ 0.0001), V (R2 ⫽ 0.67, p ⬍ 0.0001); B (R2 ⫽ 0.43; p ⬍ 0.001) and Si (R2 ⫽ 0.43; p ⬍ 0.001) in the analyzed groundwater samples. Redox conditions in the study area are oxidizing or moderately reducing. This is consistent with predominating arsenate. Thus, reductive dissolution of ferric minerals observed, for example, in Bangladesh (Smedley & Kinniburgh 2002, Ahmed et al. 2004) can be ruled out as a principal mechanism of arsenic input. In contrast, high pH values seem to promote desorption of As 63

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Figure 7. Bivariate plots of total arsenic concentration (Astot) with: (a) F⫺, (b) V, (c) B and (d) Si in shallow groundwaters. Note: the data sets with exceedingly high Astot concentrations are excluded from the plots to visualize the trends correlation.

Figure 8. (a) Distribution of Astot (␮g/L) and (b) pH in groundwater. Kriging is used to interpolate isocurves. Note irregular concentration scales.

adsorbed on to the amorphous oxides of Al, Mn and Fe. In iso-curves plots, maximum concentrations of As and high values of pH generally coincide (Fig. 8). High pH seems to be related to the dissolution of carbonates induced by cation exchange. This is consistent with a negative correlation between Ca and Na observed in earlier studies (Bundschuh et al. 2004). Another factor contributing to high pH values is probably dissolution of silicates in volcanic glass. Smedley et al. (2001) also postulated high pH as a factor contributing to high As concentrations in La Pampa region located southeast of the study area. Tentative mechanism of arsenic mobilization can be summarized as follows: (a) dissolution of volcanic ash layer with resulting release of As and Al; (b) precipitation of Al oxide and hydroxides and adsorption of As on Al and Fe mineral phases; and (c) release of As under high pH conditions. However, the importance of volcanic ash as the source of As still remains unproved and further investigation of the interaction between the ash layer and aqueous chemistry should be prioritized in further studies. 64

ACKNOWLEDGEMENTS The authors would like to acknowledge the Swedish International Development Agency (SidaSAREC) for supporting the research on arsenic-rich groundwater in the Santiago del Ester province of Argentina at the Royal Institute of Technology during 2001–2003. MC and JF acknowledge the financial support provided by the Swedish International Development Agency (Sida) in the form of Minor Field Study grants during 2002. We appreciate the constructive criticisms by an anonymous reviewer which has helped to improve the manuscript considerably.

REFERENCES Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyia, M.A.H., Imam, M.B., Khan, A.A. & Sracek, O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh, an overview. Appl. Geochem. 19(2): 181–200. Allison, J.D., Brown, D.S. & Novo-Gradac, K.J. 1991. MINTEQA2, A Geochemical Assessment Data Base and Test Cases for Environmental Systems, Athens, GA, U.S. EPA. Bejarano, G. & Nordberg, E. 2003. Mobilisation of arsenic in the Rio Dulce alluvial cone, Santiago del Estero Province, Argentina. Master Thesis, Dept. of Land and Wat. Res. Eng., KTH, Stockholm, Sweden, TRITALWR-EX-03-06, 40p. Bhattacharya, P., Jacks, G., Jana, J., Sracek, A., Gustafsson, J.P. & Chatterjee, D. 2001. Geochemistry of the Holocene alluvial sediments of Bengal Delta Plain from West Bengal, India: Implications on arsenic contamination in groundwater. In G. Jacks, P. Bhattacharya & A.A. Khan (eds). Groundwater Arsenic Contamination in the Bengal Delta Plain of Bangladesh. Proceedings of the KTH-Dhaka University Seminar, University of Dhaka, Bangladesh KTH Special Publication, TRITA-AMI REPORT 3084, pp. 21–40. Bhattacharya, P., Frisbie, S.H., Smith, E., Naidu, R., Jacks, G. & Sarkar, B. 2002a. Arsenic in the Environment: A Global Perspective. In: B. Sarkar (ed.) Handbook of Heavy Metals in the Environment Marcell Dekker Inc., New York, pp. 147–215. Bhattacharya, P., Jacks, G., Ahmed, K.M., Khan, A.A. & Routh, J. 2002b. Arsenic in groundwater of the Bengal Delta Plain aquifers in Bangladesh. Bull. Env. Cont. Toxicol. 69: 538–545. Bhattacharya, P., Welch, A.H., Ahmed, K.M., Jacks, G. & Naidu, R. 2004. Arsenic in groundwater of sedimentary aquifers. Appl. Geochem. 19(2): 163–167. Bhattacharya P., Bundschuh J., Claesson M., Fagerberg J., Storniolo A.R., Martin R.A., Thir J.M. & Sracek O. 2004. Distribution and mobility of arsenic in the Rio Dulce Alluvial Aquifer-Santiago del Estero Province, Argentina (Manuscript under submission). Bundschuh, J., Farias, B., Martin, R., Storniolo, A., Bhattacharya, P., Cortes, J., Bonorino, G. & Albouy, R. 2004. Groundwater arsenic in the Chaco-Pampean Plain, Argentina: Case study from Robles County, Santiago del Estero Province. In: P. Bhattacharya, A.H. Welch, K.M. Ahmed, G. Jacks & R. Naidu (eds) Arsenic in Groundwater of Sedimentary Aquifers, Appl. Geochem. 19(2): 231–243. Claesson, M. & Fagerberg, J. 2003. Arsenic in ground water of Santiago del Estero – Sources, mobility patterns and remediation with natural materials. Master Thesis, Dept. of Land and Wat. Res. Eng., KTH, Stockholm, Sweden, TRITA-LWR-EX-03-05, 59p. Martin, A. 1999. Hidrogeologia de la provincia de Santiago del Estero. Ediciones del Rectorado Universidad Nacional de Tucumán Argentina. Smedley, P.L. & Kinniburgh, D.G. 2002. A review of the source, behavior and distribution of arsenic in natural waters. Appl. Geochem. 17: 517–568. Smedley, P.L., Nicolli, H.B., Macdonald, D.M.J., Barros, A.J. & Tullio, J.O. 2001. Hydrochemistry of arsenic and other inorganic constituents in groundwaters from La Pampa, Argentina. Appl. Geochem. 17: 259–284. Taylor, S.R. & McLennan, S.M. 1985. The Continental Crust: Its Composition and Evolution. Blackwell Scientific, London. 365p. WHO 1993. Guidelines for Drinking Water Quality. World Health Organization, Geneva. WHO 2001. Arsenic in drinking water. Fact sheet 210: URL: http://www.who.int/mediacentre/factsheets/fs210/en/print.html (Accessed on March 9, 2004)

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Arsenic source and fate at a village drinking water supply in Mexico and its relationship to sewage contamination J.M. Cole & M.C. Ryan Department of Geology and Geophysics, University of Calgary, Calgary, Alberta, Canada

S. Smith Caminamos Juntos para Salud y Desarrollo, A.C. Cuernavaca, Morelos, México

D. Bethune Department of Geology and Geophysics, University of Calgary, Calgary, Alberta, Canada

ABSTRACT: Inhabitants of the village of Tlamacazapa display toxic health effects related to arsenic and other metal exposures from shallow well water. Arsenic is present in the dissolved phase up to a concentration of 37 ␮g/L, exceeding the World Health Organization guideline of 10 ␮g/L. Stable isotope analyses (2Hwater & 18Owater) indicate precipitation is recharging the wells through the soil and shallow groundwater. Soil and rock analyses show that arsenic is present at concentrations up to 56 mg/kg and 26 mg/kg, respectively. High concentrations of Cl, K, Na, NO3 and SO4 suggest sewage or manure contamination in the well water occurs via surface runoff into the unprotected wells or through the thin soil and shallow groundwater. Open air excretion and free-roaming animals are sources of this contamination. Arsenic and sewage contaminant concentrations are strongly correlated and the presence of sewage apparently promotes the release of arsenic from aquifer materials. It is likely that arsenic mobilization is the result of desorption associated with arsenate-phosphate competition for sorption sites.

1

INTRODUCTION

Tlamacazapa (Tlama) is located in the Sierra Madre del Sur of the state of Guerrero, roughly 160 km southwest of Mexico D.F. and 10 km east of the historical silver-mining town of Taxco de Alarcón. It has a population of approximately 9000, living in high-density housing (mixture of concrete housing and simple cornstalk/cedar branch structures) extending up a steep mountain incline. Sanitation is a critical problem due to inadequate sewage management, with most households using open-air excretion methods as opposed to constructed sanitary facilities. Symptoms of low level, chronic metal toxicity were recognized among an undernourished, inadequately hydrated and poor population by health personnel working in Tlamacazapa as part of a small, international non-governmental organization Caminamos Juntos para Salud y Desarrollo (CJ; or Walking Together for Health and Development). Hair and nail arsenic concentrations reach 23 and 290 mg/kg by dry weight, respectively, when tested on symptomatic persons admitted to hospital. Observed arsenic-related symptymology includes skin pigmentation changes (hyperpigmentation), neurological disorders and abdominal pain. Symptoms fluctuate with seasonal changes in well water levels and apparently worsen following periods of high rainfall. A multi-disciplinary research programme was initiated to study the environmental source of arsenic and other metals (especially lead), together with other contamination vectors, including soil, clay cooking pots, and palm dye used in basket weaving. As part of this multi-disciplinary research programme, the objectives of the present study are: (1) to geochemically characterize the local hydrological system, (2) to investigate the source(s) of 67

arsenic in the groundwater and (3) to assess geochemical controls on the mobility of arsenic in the aqueous environment. Water samples were collected during both wet and dry seasons for major ion, trace metal, and stable isotope analyses. Sample locations include drinking water sources for Tlamacazapa. Regional water wells and springs were sampled in order to determine the regional extent of the contamination.

2 2.1

REGIONAL AND LOCAL GEOLOGY Geological characteristics

The study area is located in the Guerrero Terrane of southern Mexico (Campa & Coney 1983). Basement rock of this terrane is composed of schists of the Tierra Caliente Complex and is overlain by Mesozoic marine and continental sequences (Fig. 1). Massive platform carbonates of the Morelos Formation and thin-bedded calcareous shales, sandstones and conglomerates of the Mexcala Formation dominate these Mesozoic sequences. Post-depositional deformation associated with the late Cretaceous Laramide Orogeny affected the Mesozoic sequence extensively and the basement sequences locally. Post-Laramide stratigraphy includes intrusive and extrusive volcanic sequences of the Tertiary Volcanic Province of Southern Mexico. These volcanics range in age from late Cretaceous to Pleistocene and are associated with a magmatic arc extending from north of the study area to the coastal regions (Morán-Zenteno et al. 1998). Thin (⬍1 m), poorly developed soils contain abundant unweathered rock fragments. Mining in Taxco de Alarcón began in pre-Hispanic times with the exploitation of silver and gold. Currently, mining is focused on the recovery of lead, zinc and to a lesser extent silver. Hydrothermal metal deposits occur as replacement veins hosted in metamorphic basement rock, Mesozoic marine

TRANS-MEXICAN VOLCANIC BELT Hiatus in Tertiary Volcanic Activity TERTIARY VOLCANIC PROVINCE OF SOUTHERN MEXICO BALSAS FORMATION Coarse continental clastics interbedded with extrusive volcanics

Regional Angular Unconformity

MEXICALA FORMATION Calcareous shales, continental sub-lithoarenites and conglomerates

CUAUTLA FORMATION Limestones and mudstones with interbedded chert

MORELOS FORMATION Platform carbonates

Regional Angular Unconformity TIERRA CALIENTE COMPLEX Roca Verde Taxco Viejo Taxco Schist

Figure 1.

Simplified stratigraphic succession in the study area.

68

sequences, and to a lesser extent in Tertiary intrusives (Salas 1991). Documented ore mineralogy includes source minerals for arsenic such as pyrite, arsenopyrite, and proustite as well as a host of other metal-bearing sulphides (Salas 1991, de Csersna & Fries 1981). As well, sulphide mineralization hosted in the Coxcatlán granodiorite about 3 km east of Tlama has been prospected and found to be economically unfeasible. Arsenian pyrite is noted to be present in this mineralization (Campa & Ramírez 1979). 2.2

Soil and bedrock geochemistry

As part of their multi-disciplinary research programme, CJ completed bulk chemical analyses of soil and bedrock from Tlamacazapa. In total, 38 composite soil samples (surface to bedrock) and two profiles were collected to provide a good geographical sample of the community as well as sites chosen based on reported metal symptoms. Bedrock samples were chosen from prominent geological formations of the Tlama area. Average soil results show consistently high concentrations of arsenic, iron, manganese and aluminum (Table 1). Soil arsenic concentrations are elevated with respect to world background soil concentrations (Boyle & Jonasson 1973). High aluminum and iron concentrations are typical of soil laterites developed in tropical climates. Bedrock results show a large variation between rock types, with elevated arsenic, copper, and strontium in the vein deposits suggesting the presence of sulphide mineralization. 2.3

Water resources

Precipitation in Tlamacazapa occurs seasonally, with arid conditions prevailing roughly from December to May. Drinking water supply is provided by up to eight shallow wells (depending on the season) excavated into existing fractures in the limestone (Fig. 2). Low concrete walls surrounding the wells limit direct surface runoff, with stairs providing access for water fetching by hand. An alternate source of water called Los Sabinos is located 5 km northeast of Tlama (Fig. 2). Water is collected and pumped uphill through three pump-and-hold stations and held in a storage tank before gravity distribution through a rudimentary piping system. This inadequate and undependable system pumps water to each neighbourhood roughly every nine (dry season) to 15 days (wet season). The water source is a shallow aquifer (⬍1 m below ground surface) located in an agricultural valley.

3

METHODOLOGY

A total of eight local, ten regional, and four surface water samples (Fig. 2) were collected at the end of both the wet (December, 2002) and dry (May, 2003) seasons for determination of field parameters (Temperature, pH, Eh, DO, EC), and analyses of major ions, trace metals and environmental isotopes (2Hwater, 18Owater, 13CDIC, 34SSO4). Trace metal samples were collected for total dissolved and total recoverable analysis, with filtered (0.45 micron) and raw water preserved with spec-pure nitric acid to a pH below 2. Arsenic analyses were done by ICP-MS (Perkin Elmer 6001). Table 1. Bulk metal analyses (in mg/kg) from different rock units (one sample each) and average soil composition (53 samples) (Cole 2004).

Limestone Granodiorite intrusion Mineralized veins Soil

As

Pb

Cu

Al

Fe

Mn

Sr

2 9 26 17

0.8 4.6 1 20

11 6 16 n/a

0.01 0.73 0.02 25000

0.03 2.21 0.11 18200

35 245 30 279

92 13 109 36

69

Figure 2. Location of study area and water sampling locations. Inset shows location and elevation of wells sampled in Tlamacazapa.

4 4.1

RESULTS AND DISCUSSION Groundwater geochemistry

Ground water is near-neutral pH (6.7 to 7.7) and mildly oxidizing (Table 2). Oxidized species include nitrate, phosphate, and sulphate, and redox potential (Eh) ranges from 170 to 220 mV. Mean water temperatures reflect average annual air temperatures of 21°C. Well water geochemistry is dominated by Ca, Mg, and HCO3, as expected for waters interacting with carbonate rocks and/or calcium-rich soils. High concentrations of Cl, K, Na, NO3-N, and SO4, and high electrical conductivity (EC) values (Table 2) indicate sewage contamination in four lowest-elevation Tlama wells (Fig. 3). The composition of the most impacted wells is comparable to septic system effluent and domestic wastewater (Table 2). Nitrate values exceed the World Health Organization (WHO) drinking water guideline (11.3 mg/L NO3-N; WHO 1998) in the four topographically lowest wells in Tlama, while the higher elevation wells have concentrations below the guideline. The presence of sewage impact in lower elevation wells (Fig. 3) suggests that Tlama is the source of its own contamination. A mixing line between uncontaminated (higher elevation) wells and sewage impacted wells is evident by increasing Na ⫹ K, Cl, and SO4 concentrations and relatively high estimated total dissolved solid concentrations (Fig. 4). In some instances, Tlama well water is more concentrated than domestic sewage (Table 2). This is likely due to the lack of dilution by grey water (e.g. toilet flushing and wash water). Wet season samples are more concentrated than dry season samples in all sewage-related parameters. The elevated wet season concentrations are likely related to increased recharge, with active sewage transport to the wells through the thin soil and shallow groundwater. 70

Table 2. Dissolved arsenic, field parameters and major ion concentrations for wet season samples from Tlamacazapa drinking water wells, Los Sabinos, and regional wells. All values in mg/L unless noted. As (␮g/L) Tlama water wells Well 1 –* Colontsintla – Well 2** 22 Well 3a** 32 Well 3b** 37 Well 4** 27 Well 5 – Mixicapan 1.8 Los Sabinos – Regional water wells Zacapalco 1 2 Taxco el – Viejo Juliantla – Coxcatlán 35 La Venta 2 Regional spring waters Zacapalco 2 – Las Grenadas 1 Cienaguillas 8 Typical –*** sewage composition

T (°C)

pH

Eh (mV)

EC DO (␮S/cm) Ca

Mg K

21 18 21 21 21 21 17 16 20

6.7 7.5 6.7 7.0 7.7 6.7 7.3 6.8 6.7

211 197 178 192 209 170 220 186 188

3 0 6 1 6 1 4 5 2

640 740 1530 1080 1040 1270 590 735 776

74 99 120 110 110 130 69 78 81

37 43 54 48 47 54 33 36 39

1 2 69 49 51 63 0 2 0.1

23 24

7.5 6.3

231 227

9 1

632 1037

77 161

17 11

22 18 21

6.6 6.9 6.7

173 181 194

1 1 2

994 1062 603

138 179 83

21 23 20 –

7.0 6.6 6.2 7.2

231 147 202 –

5 7 2 –

560 577 668 1336

58 93 91 57

Na

Cl

NO3-N PO4 SO4

HCO3

1 4 79 63 49 66 2 5 1

3 10 110 95 98 145 2 8 2

7 22 85 65 51 68 5 1 0

0.3 0.2 2 3 0.2 1 1 1 1

30 78 135 115 120 160 33 52 8

300 220 260 340 290 460 250 310 270

0.4 1

1 10

2 42

1 27

0.9 0.3

8 205

252 190

17 18 18

1 14 1

19 36 21

64 84 9

22 22 8

0.2 – 0.6

211 193 37

243 437 326

20 8 7 17

0 1 1 4 1 7 15 112

1 7 7 100

0 2 1 2.6

0.9 0.9 0 11

5 21 70 129

100 393 218 241

40

160

30

120

20

80

10

40

0 1975

Figure 3.

2025 2075 elevation (masl)

Cl (mg/L)

dissolved As (µg/L)

Note: * Results below analytical detection limit; ** Lower elevation Tlama wells; *** Data on As is not available from sources. Typical sewage composition is included (Kim et al. 2002, Robertson et al. 1991, Crites & Tchobanoglous 1998, City of Calgary unpublished data).

0 2125

Relationship of elevation with As and Cl concentrations.

The carbonate system is the dominant control on major ion concentrations in the unimpacted wells. Most waters are at or near equilibrium with calcite and dolomite, where equilibrium is defined as a saturation index [SI ⫽ Log (IAP/K)] range of ⫺0.5 to 0.5. A seasonal variation in saturation indices exists, with wet season samples showing a strong tendency towards undersaturation with respect to calcite while dry season samples are all saturated (Cole 2004). This seasonal variation is likely related to increased recharge during the wet season, resulting in dilution and less water-rock interaction. 71

Figure 4. Piper plot of Tlama wells and typical sewage concentrations (Kim et al. 2002, Robertson et al. 1991, Crites & Tchobanoglous 1998, City of Calgary unpubl. data). Circle diameter of central diamond indicates total dissolved solids. High arsenic locations indicate concentrations ⬎10 ␮g/L.

4.2

Arsenic geochemistry

Arsenic is present in five of the nine drinking water sources of Tlamacazapa. Four of the five wells used for drinking water contain arsenic concentrations exceeding the WHO drinking water guideline (10 ␮g/L; WHO 1998) and the maximum concentration is 37 ␮g/L. Although concentrations are not present at levels similar to other affected areas of the world (i.e. Bangladesh, Argentina), it is believed that detrimental health effects occur at exposure to concentrations less than 50 ␮g/L (WHO 2001). Malnutrition and dehydration are believed to compound the adverse toxic health effects. Total and dissolved arsenic results correlate closely (R2 ⫽ 0.997) suggesting that arsenic is present primarily in the dissolved phase. Arsenic concentrations are also correlated to sewage contamination (Figs. 3 and 4), with positive correlations including Cl⫺ (R2 ⫽ 0.83), NO3-N2 (R2 ⫽ 0.73), 2 K (R2 ⫽ 0.82), SO4 (R2 ⫽ 0.72) and HCO⫺ 3 (R ⫽ 0.93) (wet season, Tlama water wells). These correlations are stronger during the wet season, which is consistent with increased sewage impact during the period of higher ground water recharge. Among the sewage-related anions introduced into the Tlama wells, PO4, SO4, Cl and NO3 compete with arsenic for sorption sites (Manning & Goldberg 1996 a,b). The relative affinity for sorption to Fe (III)-oxide surfaces is PO4 ⬎ SO4 ⬎ Cl ⬇ NO3 (Rau et al. 2003). Phosphate is the best known competitor to arsenic, and displays similar geochemical behaviour to arsenic oxyanions. Phosphate concentrations decrease arsenate adsorption over a large pH range (2–11) (Manning & Goldberg 1996a, b). Arsenic mobilization by competitive sorption due to phosphate fertilizer application has been observed in agricultural soils (Peryea & Kammereck 1997). Phosphate is also known to strongly adsorb to carbonate minerals (De Kanel & Morse 1978). Although phosphate is considered the most important competitor, it is likely that all anions introduced through sewage contamination are involved as they are present at concentrations much higher than phosphate. The poor correlation of As and PO4 (R2 ⫽ 0.22) in Tlama well waters, combined with the correlation of arsenic and other sewage-derived parameters, suggests that arsenic mobilization is 72

y = 7.97x + 11.03

-50

δ2 Η ‰

-55

-60

-65

-70

-75 -11

-10

-9

-8

-7

δ18Ο ‰

Figure 5. ␦18O vs. ␦2H for Tlama well waters plotted with LMWL-Mexico City.

occurring by PO4-As competitive sorption. Phosphate is sorbed and arsenic desorbed during the infiltration of phosphate-rich sewage impacted recharge through soil and shallow groundwater. The consequent immobilization reduces its aqueous concentration with respect to the other sewage contaminants. It is also possible that faecal organics are driving the reduction of iron oxyhydroxides and the subsequent release of bound arsenic (McArthur et al. 2004). This is not believed to be the dominant process liberating arsenic because of the prevalent oxidizing conditions and lack of dissolved iron in As-rich waters. Arsenic is present at five of eight regional water wells and springs at an average concentration of 6 ␮g/L. A maximum concentration of 35 ␮g/L (Coxcatlán) in the wet season correlates with sewage contamination (Cole 2004). At these regional locations, arsenic in drinking water is a concern for lower income families, as higher socio-economic levels allow the majority of residents to purchase clean drinking water. 4.3

Stable isotopes

Stable isotope results are consistent with groundwater recharge by precipitation. Tlama well water ␦2H and ␦18O plot closely with the local meteoric water line (LMWL) for Mexico City (Cortés et al. 1997) (Fig. 5). Dry season waters may be slightly more evaporated than wet season, as would be expected.

5

CONCLUSIONS

Arsenic is present in the drinking water supply of the village of Tlamacazapa, Mexico at levels above the WHO guidelines (10 ␮g/L; WHO 1998). Average arsenic concentrations from the bedrockhosted well waters are 13 ␮g/L with a maximum concentration of 37 ␮g/L. Although arsenic is not present at levels similar to other affected areas of the world, its concentrations are significant enough to be contributing to the observed adverse toxicological effects in the undernourished residents of Tlamacazapa. Elevated levels of arsenic in the bedrock and soil of Tlamacazapa suggests that the source of arsenic contamination to the well water is geologic. Arsenic is present in well water with geochemical conditions (an oxidizing environment with neutral pH) that suggest significant immobilization by 73

sorption should occur. The correlation of arsenic contamination with sewage impact suggests that the arsenic release mechanism is anthropogenic. Sewage contamination is due to a lack of proper sanitation facilities. Sewage impact is indicated by elevated concentrations of Cl, Na, K, NO3, and SO4, and elevated EC with respect to other wells in the village. The most contaminated Tlama wells are comparable to raw sewage (Kim et al. 2002, Robertson et al. 1991, Crites & Tchobanoglous 1998, City of Calgary unpubl. data). The occurrence of sewage impact in lower elevation wells suggests that inadequate sewage management in Tlama is the source of the contamination. It is thought that the sewage contaminants are transported to wells by recharge through thin soil and shallow groundwater, particularly in the wet season. Sewage contamination introduces elevated concentrations of anions known to compete with arsenic for sorption sites, including SO4, Cl, NO3, and in particular PO4. Phosphate and arsenic have long been known to behave similarly and it is well documented that phosphate has a stronger affinity for sorbents, thus removing arsenic through competitive desorption. All sewage related parameters are correlated to arsenic in Tlama well waters, with the exception of phosphate. This suggests that phosphate is being removed by sorption while liberating sorbed arsenic.

ACKNOWLEDGEMENTS This work was supported greatly by the local volunteers working with Caminamos Juntos para Salud y Desarrollo and I. Hutcheon of the University of Calgary. Funds from the Canadian International Development Agency, the Geological Society of America, and the Central American Water Resource Network (CARA) were appreciated. We are thankful to Prosun Bhattacharya and Kazi Matin Ahmed for their comments on an earlier draft of this manuscript.

REFERENCES Boyle, R. & Jonasson, I. 1973. The geochemistry of As and its use as an indicator element in geochemical prospecting. Journal of Geochemical Exploration 2: 251–296. Campa, M. & Coney, P. 1983. Tectono-stratigraphic terranes and mineral resource distributions in Mexico. Canadian Journal of Earth Sciences 20: 1040–1051. Campa, M. & Ramírez, J. 1979. La evolución geológica y la metalogénesis del noroccidente de Guerrero. Universidad Autónoma de Guerrero. Serie Técnico-Cientifica 1: 1–84. Cole, J. 2004. Arsenic in a village drinking water supply, Mexico. University of Calgary unpublished M.Sc. thesis. University of Calgary, Calgary. Cortés, A., Durazo, J. & Farvolden, R. 1997. Studies of isotopic hydrology of the basin of Mexico and vicinity: annotated bibliography and interpretation. Journal of Hydrology 198: 346–376. Crites, R. & Tchobanoglous, G. 1998. Small and decentralized wastewater management systems. McGraw – Hill Series in Water Resources and Environmental Engineering. Boston. 1084 p. De Cserna, Z. & Fries, C. 1981. Hoja Taxco 14Q-h(7). Resumen de la geología de la Hoja Taxco, Estados de Guerrero, México y Morelos. Carta Geológica de México Serie 1:100,000. Instituto de Geología, Universidad Nacional Autónoma México. de Kanel, J. & Morse, J.W. 1978. The chemistry of orthophosphate uptake from seawater on to calcite and aragonite. Geochimica et Cosmochimica Acta 42(9): 1335–1340. INEGI 1984. Carta geologica E14-5, Hoja Cuernavaca, E 1: 250 000. Instituto Nacional de Estadística Geografía e Informática, Mexico. Kim, K., Lee, J.S., Oh, C., Hwang, G., Kim, J., Yeo, S., Kim, Y. & Park, S. 2002. Inorganic chemicals in an effluent-dominated stream as indicators for chemical reactions and streamflows. Journal of Hydrology 264: 147–156. Manning, B.A. & Goldberg, S. 1996a. Modeling arsenate competitive adsorption on kaolinite, montmorillonite and illite. Clay and Clay Minerals 44(5): 609–623. Manning, B.A. & Goldberg, S. 1996b. Modeling competitive adsorption of arsenate with phosphate and molybdate on oxide minerals. Soil Science Society of America Journal 60: 121–131.

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Morán Zenteno, D., Alba Aldaves, L., Martínez Serrano, R., Reyes Salas, M., Corona Esquivel, R. & Angeles García, S. 1998. Stratigraphy, geochemistry and tectonic significance of the Tertiary volcanic sequences of the Taxco-Quetzalapa region, southern México. Revista Mexicana de Ciencias Geológicas 15(2): 167–180. Peryea, F. & Kammereck, R. 1997. Phosphate-enhanced movement of arsenic out of lead arsenatecontaminated topsoil and through uncontaminated subsoil. Water, Air and Soil Pollution 93: 243–254. Rau, I., Gonzalo, A. & Valiente, M. 2003. Arsenic (V) adsorption by immobilized iron mediation. Modeling of the adsorption process and influence of interfering anions. Reactive and Functional Polymers 54: 85–94. Robertson, W., Cherry, J. & Sudlicky, E. 1991. Ground-water contamination from two small septic systems on sand aquifers. Ground Water 29(1): 82–92. Salas, G. 1991. Taxco Mining District, state of Guerrero. In G. Salas (ed), The Geology of North America, Economic Geology, Mexico Vol P-3: 379–380. Colorado, Geological Society of America. WHO 1998. Guidelines for Drinking-Water Quality, 2nd edition. World Health Organization, Geneva. Addendum to Volume 2. WHO 2001. Arsenic in drinking water, Fact Sheet No 2010. World Health Organization, Geneva.

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Arsenic contamination of the Salamanca aquifer system in Mexico: a risk analysis R. Rodriguez & M.A. Armienta Instituto de Geofisica, Universidad Nacional Autónoma de México (UNAM), C.U., Del Coyoacan, México City, México

J.A. Mejia Gómez Consejo Tecnico del Agua, COTAS, Irapuato-Valle de Santiago, Gto., Mexico

ABSTRACT: Solutes coming from different pollution sources are affecting the Salamanca aquifer system. The solid wastes of different chemical industries were disposed in dumps located around the urban area. Arsenic, lead and benzene are some of the reported pollutants in the local groundwater. Most of the urban wells reported As contents over the national standard for drinking water. Groundwater is the only water source for the Salamanca inhabitants. An aquifer vulnerability assessment was done using the DRASTIC method. A vulnerability zoning, a potential pollution sources inventory and a population survey permitted a risk analysis. The most critical areas are related to a subsidence fault that is acting as a preferential channel for pollutants, particularly for hydrocarbon residuals. The fault permits the hydraulic communication between an unexploited shallow aquifer and the exploited unit. Evidences of health affectation were not found. Three urban wells located near of the fault were closed.

1

INTRODUCTION

Arsenic (As) contamination in groundwater is becoming a serious problem for the institutions in charge of water supply, management and administration. In Mexico, the number of areas with groundwater As content greater than the Mexican standard for drinking water, is increasing. The associated risks of consumption of As contaminated water are well known by the authorities but not by the population. In some cases groundwater is the main source of potable water. Adverse health affects due to chronic exposure to As in several countries have been reported (Bahmra & Costa 1992, Chen & Li 1994). The drinking water standard for As in Mexico has been lowered from 0.05 mg/L stepwise by a concentration value of 0.005 mg/L/yr since 2000 in order to achieve a final standard value of 0.025 mg/L by the year 2005. This poses a significant challenge for the Mexican Health and Water Authorities to improve the analytical capacity of the laboratories. The source of As in most of the cases is natural, mainly released from the As bearing rocks and minerals. There are few reported cases in Mexico where As is anthropogenic (Armienta et al. 2000). Aquifer vulnerability assessments and water consumption patterns can help in a risk analysis. Knowledge about the risks allows taking adequate measures for efficient management of groundwater through optimizing abstraction volume as well as replacement of the wells at risk. The main objective of this study was to asses the influence of As sources location and aquifer vulnerability on the distribution of As in groundwater. The various As sources in such areas are correlated with potential health affectations. 77

2

THE SALAMANCA CASE

Salamanca City is located in Guanajuato state, central Mexico. The more than 1900 wells in the Salamanca area represent an abstraction greater than 569 Mm3/year. Groundwater is the sole source of water for the more than 140,000 Salamanca inhabitants. The main water consumer is agriculture, which accounts for more than 70% of the total abstraction. The industrial consumption is about 23%, while the domestic consumption represents approximately 7% of the abstraction. Urban supply is provided by means of 33 wells located in the urban area. The estimated deficit is 173 Mm3/year, which results in a draw down of 1.6 m/year (Guysa 1998). The Lerma river crosses the urban area, dividing it in two parts, the northern and southern areas. It received untreated urban and industrial wastewaters. The urban treatment plant is under construction. The River hydraulically divides the aquifer system. Part of the urban area of Salamanca, has been affected by subsidence. The accumulated subsidence is about 70–80 cm downtown. The subsidence process is associated with the intense abstraction of groundwater and not necessarily to aquifer over-exploitation (Holzen 1984). Similar processes have been reported in other Mexican cities (Trujillo 1991, Lesser 1998). Since 1982, a subsidence fault of more than 2 km in length has been reported resulting in damage to buildings. The fault crosses an industrial zone including a refinery and a thermal power plant. The urban area is divided in two by the fault. There is also a fault zone composed by a complex system of small faults and fractures along the main one. There are no historical records of the mean subsidence velocity rates. The first measurements indicate it is about 6 cm/year (Garduño et al. 2001). The fault affected refinery pipelines provoking an oil spill. Free phase was detected in hand made shallow wells, norias and in one of the urban wells. The refinery started a recuperation program early 1999. Arsenic and lead groundwater concentrations over Mexican standards were also reported (Rodriguez et al. 2002). Four wells have been closed due to the presence of As above the local drinking water standards, three of them in the northern area (4, 11 and 24) and one in the south (29). 2.1 The aquifer system In the northern area, the aquifer system is composed by a shallow aquifer, an intermediate unit and a deep aquifer. Fluviolacustrine sediments, fine sands, gravels and clay lenses, define the shallow aquifer. The water table depth is 18 m. The intermediate aquifer has a similar composition, with prevalence of sands and gravels and minor clay. Its static water level is located 35–40 m depth. Clay packages confine the intermediate unit. The deep aquifer is integrated by fractured Pleistocene volcanic rocks. This last formation is confined by clay layers. The piezometric level is located between 65–75 m depth (Fig. 1). The shallow unit is non-exploited. The urban wells are exploiting the intermediate formation whereas the deep unit is only used for the thermoelectric plant. The fault is hydraulically communicating the shallow aquifer with the intermediate aquifer. In the southern area there is neither shallow nor deep aquifer. There is only a non confined aquifer. The geologic environment changes drastically. To the north prevails sedimentary rocks whereas to the south Quaternary volcanic rocks domain. 2.2 Aquifer vulnerability zoning in Salamanca To assess the aquifer vulnerability, the parametric method DRASTIC (water table Depth, net Recharge, Aquifer mediates, Soil, Topography, Impact to the vadose zone and hydraulic Conductivity) was applied (Aller et al. 1985). The DRASTIC values are dimensionless. The total minimum of the Drastic Index is 21 whereas the maximum is 226. Application of this method requires reliable stratigraphic and piezometric data on the area. The vulnerability assessment was made for the shallow aquifer in the northern part of the urban area, and for the aquifers located in the southern area. The study was limited to the urban area, covering 78

Figure 1. Mexico.

Representative geological profile (N–S) across the Salamanca aquifer system, Guanajuato,

approximately 60 km2. In the vulnerability map, the fault, the green areas and the Lerma river bed were incorporated. The DRASTIC map (Fig. 2) shows that the most vulnerable area corresponds to the fault trace. The wells 4, 11 and 24 are located in this area. Low vulnerabilities are associated to the presence of clay packages of very low conductivity. The refinery lands are located on areas of very low vulnerability, but the fault, which crosses them in the southeast, makes the situation vulnerable. Medium to high vulnerability areas also exist towards south, but the main potential sources that may contain As are primarily located in the northern zone. A geo-referenced inventory of potential sources of aquifer contamination was incorporated to the vulnerability maps becoming them, maps of potential risks of aquifer pollution. Active and closed landfills, active and abandoned industrial lands, brick factories, human cemeteries, chemical contingencies, among other potential sources were incorporated.

2.3 Water consumption patterns The water consumption patterns impact the level of risk associated with the ingestion of pollutants. A survey was carried out covering a population comprising one hundred fifty families, which received water from wells that presented indicated changes in the water quality (11, 4 and 24). Results showed that the population has a particular perception on the problem of the water contamination. In spite of the fact that the population identified scent, flavor and color related with the presence of hydrocarbons, only 48% consumes bottled water regularly, while 16% is only supplied with municipal water. The cost of the purified water also impacts in its relatively low consumption. The price of the supplied water is approximately US$ 0.30 per cubic meter, whereas the price of purified water is US$ 2.00 per 20 liters. In Mexico it is common that the population has underground storages for water, cisterns, and tanks in the superior part of the houses, “tinacos”. These temporary containers avoid direct consumption of the water distributed by the pipeline network for drinking purposes. In Salamanca 30% of the population that received water of the wells, 4, 11 and 24 does not have cistern or tinaco, which implies that they received water directly from the pipeline net. Direct consumption increases the exposure risk to volatiles. Only 2% of the total population has some formal system of filtration of water. The exposure period is an important factor for potential adverse health 79

Figure 2. Map of Salamanca showing the risks to groundwater contamination based on a DRASTIC vulnerability assessment.

effects. Sixty five percent of the population have lived in the same house, for more than 10 years, with the same pattern water consumption. However, correlations between cancer incidence and any risk factor were not found.

3

ARSENIC DISTRIBUTION

Arsenic has been detected in almost all the urban wells in Salamanca area at concentration levels between 0.01 to 0.05 mg/L. The highest concentration was 0.28 mg/L in well 29, during late 2003. The concentration distribution for May 2004 is shown in Figure 3. Maximum concentration of As 80

Figure 3.

Variation of As concentration in the Salamanca aquifer system during May 2004.

were measured in the southern area (0.10–0.05 mg/L), but in the NW area one well indicated high As concentration. The As content in the shallow aquifer is lower than the As detected in the intermediate unit (Rodriguez et al. 2002). The highest As concentration during 2003 was located in the fault area whereas for 2004 the maxima are located in the south (Fig. 3). Wells 4, 11 and 24 are located in the vicinity of the fault, where highest vulnerability values were found, were closed. This area is supplied by wells located to the west, outside the fault area. Before the well closure, the As maximum was located in that zone. The local hydrodynamics changed after the well closure. It could have influence over the local hydrogeochemistry.

4

CONCLUSIONS

The location of active contamination sources in high aquifer vulnerability zones may result on groundwater quality problems due to the dissolution of solutes from the sources and its later leakages toward the aquifer. Abstraction and distribution of polluted water becomes a health risk to the supplied population. The health risk maintains a linear dependency with the water consumption patterns. The fault trace defines one of the most vulnerable areas (Fig. 2). Hydrocarbons free phase in one of the urban wells, and chloroform in one of the piezometers, may have resulted from contaminant mobilization through this fault. Chloroform may originate from the reaction of organochloride compounds leaked from the factories, with residual chlorine coming from leakages of the drinking water pipelines; sewage infiltration along the fault, could also originate chloroform through the same process. During the operation of the most polluted well, high As concentrations were also detected in that area. Various authors have found correlation between vulnerable areas and aquifer contamination (Kalinsky 1994, Báez 2001, Ramos 2001). The residual hydrocarbons are able to move along the fault and even to continue contaminating the intermediate aquifer, which is the only source of drinking water. The thickness of the free 81

phase presented strong variations, of centimeters to almost two meters in 24 hours, indicating that there are factors that mobilize the hydrocarbons, as changes in the abstraction regime. It is not discarded the possibility of another active source. Presence of clay lenses and paleochannels may influence the migration of contaminants. High permeable paleochannels, found during urban excavations, may act as preferential conduits for polluted water. The risk level is increased by the presence of various pollutants and by their probable combined action. Two carcinogens of different origin, As and organic compounds, acting on the same organ or tissue may increase their damage. There is no local epidemiological information that may allow the determination of the degree of correlation between As exposure and cancer, although this possibility exists. There are crossed exposure routes from different As sources. Groups of people have consumed water with concentrations of arsenic above the drinking water standards and live in the vicinity of the industrial areas, being exposed to powders and emanations. Workers exposed to As in their working places also live in the areas that were supplied with water from the three polluted wells. The extension of the fault is intimately linked with the affectation to the aquifer system, and consequently with the health risk of the supplied population. Areas of very low vulnerability determined in early 2000, may change their status because the extension of the fault can facilitate the solute leakage towards the aquifer.

ACKNOWLEDGEMENTS This project was financed by the grant 003 of the National Council of Science and Technology, (CONACyT,-SEMARNAT) Mexico. The authors thank CMAPAS (Municipal Council of Potable Water of Salamanca) the facilities of access to the information. Chemical analyses were carried out at the Analytical Chemical Laboratory at the IGF by O. Cruz, A. Aguayo and N. Ceniceros.

REFERENCES Armienta, A., Rodriguez, R., Morton, O., Cruz, O., Ceniceros, N., Aguayo, A. & Brust, H. 2000. Health risk and sources of arsenic in the potable water of a mining area. In E.G. Reichard, F.S. Hauchmann & A.M. Sancha (eds), Interdisciplinary Perspectives on Drinking Water Risk Assessment and Management, Proc. of the Santiago (Chile) Symposium, September 1998. IAHS Publication 260: 9–16. Aller, L.T., Bennet, J.H., Lehr, R., Petty, J. & Hackett, G. 1985. DRASTIC; A Standard System for Evaluation Groundwater Pollution using Hydrogeologic Setting, Publication EPA/600/2-85/081, US EPA, 622 pp. Báez, A. 2001. Verificación y validación del Indice AVI de León Gto. Tesis Maestría en Protección y Conservación Ambiental, Univ. Iberoamericana campus León, Mexico, 90 pp. Bahmra, R.K. & Costa, M. 1992. Trace elements Aluminium, Arsenic, Cadmium, Mercury and Nickel. In Environmental Toxicants. Ed. Lippmann M. Van Nostram Edit., New York US, 575–632 pp. Chen, C.J. & Lin, L. 1994. Human carcinogenicity and atherogenicity induced by chronic exposure to inorganic arsenic. In Arsenic in the Environment, Part II: Human Health and Ecosystem Effects (ed. by O. Nriagu) John Wiley and Son Edit., New York, 109–131 pp. Garduño, V.H., Arreygue, E., Hernández, V.M., Rodriguez, G.M. & Valencia, E.O. 2001. Estudio sobre los riesgos geológicos en el Municipio de Salamanca Gto. (Study of the geological risks in Salamanca City) Technical Report. Univ. Michoacana de San Nicolás de Hidalgo, Municipio de Salamanca, Mexico, 36 pp. GUYSA 1998. Estudio Hidrogeológico y Modelo matemático del Acuífero del Valle de Irapuato – Valle de Santiago. CEASG, Guanajuato, Mexico. Technical Report APA-GTO-97-023, 230 pp. Holzen, T.L. 1984. Ground failure induced by groundwater withdrawal from unconsolidated sediments. Man induced land subsidence. Geological Soc. of America, Reviews in Engineering Geology VI: 67–105. Kalinski, R.J., Kelly, W.E., Bogardi, I., Ehrman, R.L. & Yamamoto, P.D. 1994. Correlation between DRASTIC vulnerabilities and incidents of VOC contamination of Municipal Wells in Nebraska. Groundwater 32(1): 31–34. Lesser, J.M. & Cortes, M.A. 1998. El hundimiento del terreno en la ciudad de Mexico y sus implicaciones en el sistema de drenaje. Ing. Hidráulica en México 3: 13–18 (In Spanish).

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Ramos, A. 2001. Validación de mapas de vulnerabilidad e impacto ambiental. Caso Río Turbio. PhD Thesis Doctorado en Aguas Subterráneas, Posgrado en Ciencias de la Tierra UNAM, México, 98 pp. (In Spanish). Rodríguez, R., Armienta, M.A., Berlin, J. & Mejia, J.A. 2002. Arsenic and lead pollution of the Salamanca aquifer, Mexico: origin, mobilization and restoration alternatives. In S.F. Thornton & S.E. Oswald (eds): Groundwater Quality: Groundwater Quality: Natural and Enhanced Restoration of Groundwater Pollution, IAHS Publication 275: 561–565. Rodríguez, R., Ramos, A. & Armienta, M.A. 2003. Groundwater arsenic variations: The role of local geology and rainfall. Applied Geochemistry 19(2): 245–250. Rodríguez, R., Rodríguez, L. & Palma, F. 2002. Social Perception of polluted water consumption risk. An approximation between aquifer vulnerability assessment and water supply management in Salamanca Mexico. In C.A. Brebbia (ed), Risk Analysis III: 469–474, Southampton, WIT Press. Trujillo, C.J.A. 1991. Fallamiento de terrenos por efecto de la sobreexplotación de acuíferos en Celaya, Guanajuato, México. Proceedings XXIII IAH Congress, Aquifer Overexploitation, Tenerife Spain, pp. 175–178 (In Spanish).

83

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Arsenic pollution in aquifers located within limestone areas of Ogun State, Nigeria A.M. Gbadebo Department of Environmental Management and Toxicology, College of Environmental Management Resources, University of Agriculture, Abeokuta, Ogun State, Nigeria

ABSTRACT: This paper presents the results of physico-chemical analysis of groundwater samples obtained from twenty (20) hand-dug wells located in ten communities within the limestone belt of Ogun State with a view to determine the quality (or portability) in terms of arsenic concentration of the water resources of the aquifers in these communities. The result indicates a minimum of 0.04 mg/L obtained in the wells of Akinbo and a maximum of 0.16 mg/L obtained in the wells of Lapeleke. The mean values of arsenic in all the sampled aquifers are generally higher than the maximum permissible value of 0.01 mg/L recommended for drinking water by WHO. Also, the main arsenic concentration of soil, shale and limestone in the area are 22.3 ␮g/g, 78.0 ␮g/g and 514.0 ␮g/g respectively while the pollution index (PI) for the water in the aquifers generally for the limestone area is 8.1. Thus both the measured values of this metal and the calculated PI inferred groundwater arsenic pollution in this sedimentary terrain.

1

INTRODUCTION

Aquifers are the major source of clean water all over the world. However, the advent of industrial revolution and mechanized farming has drastically changed the quality of groundwater resources by introducing a wide range of chemical contaminants (Nickolas 1996). Besides, widespread problem of aquifer contamination by chlorinated solvent emanating from dry-cleaning industry together with hydrocarbon leakages from buried tanks are also sources of aquifer pollution. All these have deteriorated and degraded the quality of groundwater. Aquifer pollution may arise from the chemistry of the host rock thereby changing the chemical constituents of the interstitial water. A sand dominated aquifer is also more vulnerable to the influence of human activities than the aquifer located between two layers of impermeable materials. Thus geology plays a prominent role in determining the quality of underground water resources. Among the widely reported inorganic pollutants in groundwater, arsenic (As) is of major concern all over the world. Chronic arsenic poisoning has been reported in many parts of the world such as Bangladesh, India, Taiwan, Argentina and Chile due to the prolonged ingestion of polluted groundwater (Bhattacharya et al. 2002, Armienta 2003). According to Bhattacharjee et al. (2003) Bangladesh is the worst grip of mass poisoning the world has ever witnessed. Groundwater arsenic enrichment has been ascribed to either mineralization or geochemical mobilization. (Armienta et al. 1997, 2001) while Gonzalez-Hita et al. (1991) and Cebrian et al. (1994) have traced the sources of arsenic groundwater pollution in some rural areas to hydrothermal processes. Further, arsenic pollution in aquifer has also been ascribed to anthropogenic sources (Gutierrez et al. 1996). This study is carried out in open dug wells from selected communities located within the limestone belt where quarrying activities are currently in progress. The purpose of this study is to determine the arsenic concentration level in the well water widely consumed by the residents of these communities and at the same time highlights its health implications. This work is a follow up of research to a previous study in which high levels of arsenic were obtained in the soils collected 85

Figure 1.

Map of Ogun State showing the geological formations and well location.

from these areas. So far, there is no reported case or observable symptom of endemic arsenic poisoning in the study area. However, the findings are meant to draw the attention of the concern government on this issue.

2

LOCATION AND GEOLOGICAL SETTING

The study area (i.e. Ewekoro Limestone belt of the Dahomey Basin) falls within latitude 6°50⬘N and 7°15⬘N and longitude 3°05⬘E and 3°20⬘E. The major communities in which the well water were collected in the area include: Akinbo, Elebute, Itori, Otun, Lapeleke, Egbado, Papalanto, Ilaro, Aiyetoro, Igbogila (Fig. 1). The topography is divided into two by the southeast trending depression in which Ewekoro Formation is deposited. The outcrop area is slightly above 420 m above the mean sea level (Tijani & Ayodeji 2001). The drainage is dense and the major rivers include Ogun, Igbin, Ewekoro, Yewa, Atuwara, Akinbo etc. The area falls within the tropical rainforest zone of Nigeria with annual rainfall ranging between 1500 to 2000 mm and mean annual temperature ranging between 27 and 35°C with two distinct dry and wet seasons (Iloeje 1991, Akanni 1992, Oguntoyinbo et al. 1983). Geologically, the area of the present study forms a part of the Dahomey Basin which is one of the sedimentary basins on the continental margin of the Gulf of Guinea extending from the southeastern Ghana in the west to the western flank of the Niger Delta (Reyment 1965, Jones & Hockey 1964). The Dahomey Basin was formed consequent to the opening of the South-Atlantic during the Neocomian. The stratigraphy and the hydrogeology of this basin (Table 1) has been widely studied by various authors (Jones & Hockey 1964, Fayose 1970, 86

Table 1.

Stratigraphic units of the Dahomey Basin (Tijani & Ayodeji 2001).

Jones & Hockey (1964)

Fayose (1970)

Omatshola & Adegoke (1981)

Age Recent

Formation Alluvium

Age Early Miocene

Formation Benin

Pleistocene to Oligocene

Coastal Plain sands

to Late Oligocene

Ogwashi Asaba

Late Middle & Early Eocene Paleocene

Ilaro

Late Senonian

Ewekoro

Early Oligocene To Middle Eocene Early Eocene and Paleocene

Age Pleistocene to Oligocene

Formation Coastal Plain sands

Eocene

Ilaro Ososun

Ameki Akinbo Paleocene Ewekoro Araromi Formation

Imo Maastrichtian

Abeokuta

Late Cretaceous Upper Maastrichtian

Afowo Formation

Abeokuta to Neocomian

Precambrian

Crystalline

Crystalline Basement

Ise Formation

Omatsola & Adegoke 1981, Idowu et al. 1999). However, the sedimentary rock of the study areas in which the aquifers are located are essentially basal conglomerate and arkosic sandstone of Abeokuta Formation; the limestone and shale members of Ewekoro Formation; the massive and poorly consolidated sandstone of Ilaro Formation followed by poorly sorted coastal plain sands and alluvium. Also, the underlining crystalline basement complex rock types in the study area comprises of gneises, older granites, pegmatites etc.

3

METHODS

A total of 20 dug well water samples from ten different communities in the study area were collected and analyzed. Total depth of the wells, static water levels, temperature, pH and conductivity were determined in the field while the arsenic concentration was determined in the laboratory. The pH of the water samples were measured using field pH meter and the conductivity was measured using WTWLF 95 conductivity meter. The temperature was measure using celcius thermometer. Further, twelve soil and rock samples were randomly collected from the study area and also analyzed for arsenic concentration using Atomic Absorption Spectrophotometer (Buck 200A model). Statistical analysis using regression method and various graphs were plotted to observe the correlation pattern of the arsenic concentration with different parameters. The use of Pollution Index (PI) was adapted to predict the level of hazard associated with the aquifers.

4

RESULTS AND DISCUSSION

The chemical data on the water, rock and soils samples are shown in Tables 2 and 3. It is observed that Lapeleke has the highest arsenic concentration with mean value of 0.16 mg/L followed by Elebute and Aiyetoro with mean value of 0.10 mg/L. The least value of 0.04 mg/L was obtained in the water from aquifer located in Akinbo. These values are higher than that recommended) for drinking water by World Health Organization (0.01 mg/L; WHO 1999 and 2001). 87

Table 2. Range and mean values of physico-chemical characteristics and arsenic concentrations in groundwaters. Sample location**

Temperature (°C)

pH

Depth (m)

Electrical conductivity (␮S/cm)

Akinbo Elebute Egbado Lapeleke Itori Otun Papalanto Ilaro Igbogila Ayetoro *WHO *FEPA

24.0–24.2 23.0–23.5 23.0–24.0 24.0–24.1 23.8–24.2 23.5–23.8 23.9–24.1 24.1–24.2 23.7–25.0 24.1–24.8 25

7.1–7.3 7.0–7.5 7.4–7.8 7.1–8.0 6.9–7.2 7.0–7.1 7.2–7.6 6.8–7.0 6.9–7.0 7.0–7.3 6.0–8.5

5.4–10.3 3.6–4.2 3.6–4.2 1.9–2.3 6.8–8.8 11.6–12.9 3.5–3.7 8.2–8.5 9.1–9.4 8.8–9.2

487–894 316–334 146–193 151–532 399–874 41–332 388–392 350–850 402–905 405–875

As conc. (mg/L) 0.00–0.08 0.06–0.14 0.06–0.08 0.09–0.22 0.03–0.09 0.03–0.06 0.00–0.12 0.05–0.20 0.08–0.19 0.07–0.13 0.01 0.01

1000

Mean As conc. (mg/L) 0.04 0.10 0.07 0.16 0.06 0.05 0.06 0.08 0.09 0.10

PI (Aquifers)

8.1

* Drinking water standards. ** Duplicate samples were collected at each site.

Table 3.

Arsenic concentration in soil and the lithological units in the study area.

Sample type

Sample no.

Range of As concentration (mg/kg)

Mean of As concentration (mg/kg)

Soil Shale Limestone

4 4 4

5.2–39.5 77.7–78.5 478.0–514.2

22.3 78.0 514.0

As (mg/L)

0,20 0,15 0,10 0,05 0,00 23,1

23,4

23,7

24

24,3

24,6

Groundwater temperature (°C)

Figure 2.

Relationship between As concentration levels with the groundwater temperatures in the wells.

There seems to be a relationship between arsenic concentration and temperature of the aquifer (Fig. 2). In general, at temperatures ⬎24°C, arsenic mobility into the water is greater, while the very low value of arsenic concentration in the aquifers of Akinbo community at temperature of 24.1°C may be attributed to either the low mineralization in the shale and limestone lithologic units or it’s low mobility in the aquifers in this area. According to Smith et al. (1998), industrial effluents and geological formations are the major sources contributing arsenic to groundwater. The maximum pH value of the analyzed groundwater ranged between 6.9–7.6. However, the values fall within the WHO (1999) recommended pH value range of 6.5–8.5. These values conform well with the pH values of the arsenic polluted groundwater also from the shallow aquifers 88

As (mg/L)

0,20 0,15 0,10 0,05 0,00 6,8

7

7,2

7,4

7,6

7,8

Groundwater pH

Figure 3.

Relationship between arsenic concentration and groundwater pH.

As (mg/L)

0,20 0,15 0,10 0,05 0,00 0

2

4

6

8

10

12

14

Aquifer depth

Figure 4. aquifers.

Bivariate plot showing the variation of the groundwater As concentration with the depth of the

in the La Banda County, Argentina which Bejarano et al. (2003) described as generally near neutral to alkaline (pH 6.67–8.95). Bethany & Jankowaki (2003) opined that arsenic may be attached to mineral surfaces but are subsequently released due to changes in redox conditions with other oxyanions e.g. oxides of Fe, Mn and sulphide. Comparison of range values of pH and range values of arsenic in Table 2 indicates that there is tendency that at higher pH, more arsenic may be released. There is a little deviation from this pattern in the mean – mean plot of pH and arsenic respectively (Fig. 3). This deviation is probably due to the abundance of bicarbonate materials in the limestone components of the aquifers which Bethany & Jankowski (2003) has noted to be another member which can scavenge arsenic. Probably, being a sedimentary terrain and a limestone belt, the maximum mean depth of aquifer in the study area is 12.3 m while the least mean depth is 3.62 m. There is seemingly an inverse correlation between aquifer depth and arsenic concentration in the study area. A downward but not perfect progressive decrease in arsenic concentration was observed as the depth of aquifer increases in the study area (Fig. 4). At an aquifer mean depth of 2.12 m arsenic concentration of 0.16 mg/L was obtained while at a mean depth range of 7.79–7.85 m the mean arsenic concentration of 0.06–0.04 mg/L was obtained. Also at an aquifer depth of 12.27 m the arsenic mean concentration recorded is 0.03 mg/L. This trending pattern in aquifer arsenic concentration has been observed by Welch et al. (2003) in Bangladesh where it was discovered that arsenic concentration exceed 100 mg/L at depth of between 23 and 30 m below land surface while the arsenic concentration were ⬍5 mg/L in the deep aquifer. Also, Wagner et al. (2003) noticed that arsenic content of sedimentary aquifer in West Bengal, India decrease with the increasing depth. This was attributed to the vertical transport of arsenic within the aquifer. The electrical conductivity in the aquifer ranges between 137 and 686 ␮S/cm, which is equivalent of the Calculated Total Dissolved Solids (CTDS) range of between 82.2 and 401.6 mg/L respectively when multiplied by a factor of 0.6. Thus the EC values are indications of the presence of dissolved solutes, hence also the pollutants in groundwaters. It is evident that arsenic is one of the major contaminants present in groundwater of all the aquifers. There appears to be an undulating kind of relationship between the electrical conductivity and the arsenic concentration in groundwaters of the different aquifers (Fig. 5). 89

As (mg/L)

0,20 0,15 0,10 0,05 0,00 0

100

200

300

400

500

600

700

800

Electrical conductivity (␮S/cm)

Figure 5. Table 4.

Asconc AsconclogTf

Relationship between As concentration and electrical conductivity of the groundwater. Summary of values of regression analysis obtained for Asconc and AsconclogTf for the study areas. Constant

Temp. (°C)

pH

Depth (m)

EC (␮S/cm)

R2

⫺0.653 (⫺0.543) ⫺8.729 (⫺0.620)

0.105 (1.932) 1.173 (1.841)

⫺0.218 (⫺2.001) ⫺2.696 (⫺2.006)

⫺1.882 (⫺2.308) ⫺0.207 (⫺2.429)

⫺1.895 (⫺1.544) ⫺2.214 (⫺1.543)

57.5 58.2

Values in the parenthesis are the t-values.

The calcium content of the underground water from the aquifers ranged from 4.87 to 30.45 mg/L while the magnesium is from 0.83 to 10.24 mg/L. Both sodium and potassium constituents ranged from 14.35 to 56.66 mg/L and 1.02 to 44.59 mg/L respectively. Similarly, chloride mean concentration range of 11.57 to 85.08 mg/L and a bicarbonate range of 77.79 to 264.00 mg/L were obtained. Both the sulphate and nitrate ions mean concentrations ranged from 2.92 to 50.79 mg/L and 10.16 to 44.36 mg/L respectively. The chemical data revealed a similarity in the concentration of both the cation and anion constituents of the aquifers with the magnitude of concentration decreasing in the order of Na ⬎ K ⬎ Ca ⬎ Mg for the cations and HCO3 ⬎ Cl ⬎ SO4 ⬎ NO3 for the anions. This sequence according to Olatunji et al. (2001) is an indication of the hydrolytic processes that has taken place between the water in the aquifer and the mineral components of the hosting bedrock under confined pressure and temperature. Also, the high concentrations of the ions in the underground water according to Tijani & Ayodeji (2001) may be related to the ionic affinity among the cations and the anions since sodium occurs with potassium and calcium occurs with magnesium in the aquifers while calcium has affinity for carbonate or bicarbonate. On the basis of their chemical composition the water characterization in the aquifer of the study area fit into the three main types, viz. Na-(K)-Cl-SO4, Ca-(Mg)-Na-(K)-SO4 and Ca-(Mg)-Na-(K)-HCO3 das reported by the earlier works carried out by Tijani & Ayodeji (2001) on the surface and groundwater of the Dahomey basin. Apart from arsenic the value of nitrate (10.16 to 34.36 mg/L) was found to be above the WHO (1980) maximum recommended value of 10 mg/L while other chemical compositions of the aquifers are within the recommended limits. This implies that both arsenic and nitrates are the major pollution sources of the aquifers in the study area. The pollution index (i.e. the ratio of arsenic concentration to the recommended standard values) calculated for the groundwater geologic resource in all the aquifers from the entire limestone belt in the study area far exceeds 1 (i.e. PI ⬎ 1). This according to Kim et al. (1978) further confirms the anthropogenic source in addition to the natural source of arsenic in the study area. The result of the regression analysis carried out on the measured parameters in all the wells from the study area (Table 4) indicate that there is generally a negative statistical relationship between the arsenic concentration (Asconc), transformed log of arsenic concentration (AsconclogTf) and other parameters. This negative relationship is statistically significant for the arsenic concentration, transformed log of arsenic concentration, depth and pH (i.e. t ⬎ 2.0) but statistically insignificant with the 90

electrical conductivity (t ⬍ 2.0). However, there exist a positive and statistically insignificant relationship between the arsenic concentration, transform log of arsenic concentration and temperature in the study area. In the aquifers of the study areas, a total of 57.5% and 58.2% of the observed variation in the arsenic concentration and transformed log of the arsenic concentration respectively were accounted for by both the temperature, pH, depth, electrical conductivity while the remaining 41.8–42.5% are likely to be accounted for by geology of the aquifer, infilteration and percolation effect, atmospheric deposition etc. The statistically significant relationship between the arsenic concentration, depth and pH and also very high percentage of the observed variation in the transformed log of the arsenic concentration could be attributed to the high content of arsenic in the limestone unit in which the aquifers are located. Thus, the deeper the aquifers penetrate into the limestone units, the higher the arsenic concentration.

5

CONCLUSIONS

Occurrence of arsenic has been found in groundwater from all aquifers located within the limestone area of Ogun state under a neutral pH of 6.8–7.5. The near neutral pH favours the occurrence of arsenic in the groundwater environment (see viz. Bhumbla & Keefer 1994, Bhattacharya et al. 2002) in two main inorganic forms: As(III) and As(V). These two forms of inorganic arsenic are mostly found in natural and drinking water (Marisol Vega et al. 2003). The reduced form As(III) is more mobile and more toxic than the oxidized form As(v). In this study, the total arsenic [As(III) ⫹ As(V)] were determined. It is therefore necessary to reduce the carcinogenic effects of arsenic to the recommended standard by chemical oxidation methods using chloride, chloride dioxide and potassium permanganate (Welte 2002) or bacteria (CAsO1 – genus Thiomonas and genus Ralstonia) oxidation under authotrophic conditions in a pH range of 4 to 8 with a pH optimum of 6 (Battaglia-Brunet et al. 2002).

REFERENCES Akanni, C.O. 1992. Climate. In S.O. Onakomaiya, O.O. Oyesiku & F.J. Jegede, (eds), Ogun State in Maps. 207p. Ibadan: Rex Charles Publication. Armienta, M.A. 2003. Arsenic Groundwater Pollution in Mexico. Medical Geology Newsletter 6: 4–6. Armienta, M.A, Rodriguez, R., Aguayo, A., Ceniceros, N., Cillasenor, G. & Cruz, O. 1997. Arsenic Contamination of groundwater at Zimapán, Mexico. Hydrogeology Journal 5: 39–40. Armienta, M.A., Villasenor, G., Rodriguez, R, Ongley, L.K. & Mango, H. 2001. The role of arsenic–bearing rocks in groundwater pollution at Zimapan Valley, Mexico. Environmental Geology 40: 571–581. Battaglia-Brunet, F., Dictor, M.C., Garrido, F., Crouzet, C., Morin, D., Dekeyser, K., Clarens, M. & Baranger, P. 2002. An arsenic(III)-oxidizing bacterial population: selection, characterization and performance in reactors. J. Applied Microbiology 93: 656–667. Bejarano, G., Nordberg, E., Bhattacharya, P., Martin, R.A., Storniolo, A.R. & Bundschuh, J. 2003. Groundwater Arsenic in the shallow Alluvial Aquifers of La Banda County in Santiago del Estero Province, Argentina. Proc. 7th International Conference on Biogeochemistry of Trace Elements, Uppsala, Sweden, Vol. 2: 12–13. O’Shea, B. & Jankowski, J. 2003. The use of solid phase selective extraction techniques to support groundwater chemical data from a coastal aquifer affected by elevated arsenic concentrations. Proc. 7th International Conference on Biogeochemistry of Trace Elements, Uppsala, Sweden Vol. 2: 43–44. Bhattacharjee, M., Sultana, S., Hasneen, A., Alauddin, M., Fasconaro, M., Alauddin, S., Hussam, A. & Sikder, A.M. 2003. Speciation of Arsenic In Bore-Hole Sediment Leachate and Groundwater of Bangladesh. Proc. 7th International Conference on Biogeochemistry of Trace Elements, Uppsala, Sweden Vol. 2: 56–57. Bhattacharya, P., Jacks, G., Frisbie, S.H., Smith, E., Naidu, R. & Sarkar, B. 2002. Arsenic in the Environment: A Global Perspective. In B. Sarkar (ed), Heavy Metals in the Environment: 147–215. New York, Marcel Dekker. Bhumbla, D.K. & Keefer, R.F. 1994. Arsenic Mobilization and Bioavailability in Soils. In J.O Nriagu (ed.) Arsenic in the Environment Part 1. Cycling and Characterization, 51–81, New York, JohnWiley & Sons Inc.

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Cebrian, M.E., Albores, A., Garcia-Vargas, G., Del Razo, L.M. & Ostrosky-Wegman, P. 1994. Chronic Arsenic poisoning in humans: the case of Mexico. In J.O Nriagu (ed.) Arsenic in the Environment Part II, pp 93–107. Fayose, E.A. 1970. Stratigraphy and Paleontology of the Afowo Well. Jour. Nig. Min. Geol. Metal Soc. 5: 1–99. Gilligan, L.B, Brownlow, J.W. & Henley, H.F. 1992. Dorrigo-Coffs Harbour 1:250,000 Metallogenic Map SH/56–10, SH/56–11: Metallogenic Study and Mineral Deposit Data Sheets. New South Wales Geological Survey, Sydney. Gonzalez-Hita, L., Sanchez, L. & Mata, I. 1991. Estudio hidrogeoquimico e isotopico del acuifero granular de la Comerce Lagunera. Institute Mexicano de Technologia del Aqua, Morelos, Mexico. Gutierrez, P.A, Rodriguez, R.E, Romero, G & Velazquez, G.A. 1996. Eliminacion de arsenico en aqua potable de pozos. Actas INAGEQ 2, 319–332 Idowu, A.O., Ajayi, O. & Martins, O. 1999. Occurrence of groundwater in parts of the Dahomey Basin, Southwestern Nigeria. Jour. Min. & Geol. 35 (2): 229–236. Iloeje, N.P. 1981. A New Geography of Nigeria. Longman, Nigeria, 201 p. Jones, H.A. & Hockey, R.D. 1964. The Geology of part of southwestern Nigeria. Geol. Surv. Nigeria Bull. 31, 101p. Kim, K.W., Lee, H.K. & Yoo, B.O. 1998. The environmental impact of gold mine in the Ygu-Kwangcheon Au–Ag metallogenic province, Republic of Korea. Env. Tech. 19: 291–298. Vega, M., Carretero, C., Elices, B. Barrado, E. & Pardo, R. 2003. Flow analysis of inorganic species in groundwater by stripping voltammetry at tubular gold electrodes. Proc. 7th International Conference on Biogeochemistry of Trace Elements, Uppsala, Sweden Vol. 4: 98–99. Nickson, R.T., Mc Arthur, J. M., Burgess, W.G., Ravenscroft, P. Ahmed, K.M. & Rahman, M. 1998. Arsenic Poisoning of Bangladesh Groundwater. Nature 395: 398. Oguntoyinbo, J.S., Areola, O.O. & Filani, M. 1983. A Geography of Nigeria Development (2nd Ed.). Heinemann Educational Books Nig. Ltd. 450p. Olatunji, A.S., Tijani, M.N., Abimbola, A.F. & Oteri, A.U. 2001. Hydrogeochemical Evaluation of the Water Resources of Oke-Agbe Akoko, South-Western Nigeria. Water Resources Jour. of NAH 12: 88–93. Omatsola, M.E. & Adegoke, O.S. 1981. Tectonic Evolution and Cretaceous Stratigraphy of the Dahomey Basin. Jour. Min. and Geology Nigeria 18(1): 130–137. Rayment, C.A. 1965. Aspects of the Geology of Nigeria. Univ. of Ibadan Press, Ibadan. Smith, E., Naidu, R. & Alston, A.M. 1998. Arsenic in the soil environment: A review. Advances in Agronomy 64: 149–194. Tijani, M.N. & Ayodeji, O.A. 2001. Hydrogeochemical assessment of surface and groundwater resources in part of Dahomey Basin, South Western Nigeria. Water Resources Jour. of NAH 12: 88–93. Wagner, F., Berner, Z., Stüben, D. & Agarval, P. 2003. On the Mechanisms of Mobilization and Transport of Arsenic in Groundwater and their Consequences for Cultivated Plants In West Bengal, India. Proc. 7th International Conference on Biogeochemistry of Trace Elements, Uppsala, Sweden Vol. 2: 34–35. Welch, A.H., Stollenwerk, H.G., Breit, G.N., Foster, A., Yount, J.C., Whitney, J.W., Uddin, M.N., Alam, M.M. & Islam, M.S. 2003. Arsenic transport in groundwater of Bangladesh: Implications for use of oxidized PreHolocene aquifers. Proc. 7th International Conference on Biogeochemistry of Trace Elements, Uppsala, Sweden Vol. 2: 2–3. Welte, B. 1996. L’arsenic Techniques Sciences Methodes 5: 36–45. World Health Organization (WHO) 1993. Guidelines for drinking–water quality, 2nd Ed., Vol. 1, Recommendations. Geneva. pp 41–42. World Health Organization (WHO) 2001. Arsenic in drinking water. Fact sheet 210: URL: http://www.who.int/mediacentre/factsheets/fs210/en/print.html (Accessed on March 9, 2004)

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Section 2: Environmental health assessment-arsenic in the food chain

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Arsenic in groundwater and contamination of the food chain: Bangladesh scenario S.M. Imamul Huq Department of Soil, Water & Environment, University of Dhaka, Dhaka, Bangladesh

Ravi Naidu Centre for Environmental Risk Assessment and Remediation (CERAR), University of South Australia, Mawson Lake Campus, Adelaide, Australia

ABSTRACT: Ingestion of arsenic (As) contaminated groundwater is the major cause of As poisoning in Bangladesh. However, poisoning among the population is not consistent with the level of water intake. Moreover, there is also a spatial variation of the manifestation of arsenicosis in the country. This has raised the question about the role of the food habit, nature and amount of food intake in the As dilemma. Even if an As-safe drinking water supply is ensured yet, the same As contaminated groundwater will continue to be the main source of irrigation for about 40% of the net cultivable area as more than 60% of irrigation water comes from groundwater. This leaves a risk of soil accumulation of the toxic element and eventual exposure of the food chain to As contamination through plant uptake and animal consumption. Water (hand tube wells and irrigation pumps), soil and vegetables/crop samples collected from as many as 150 different locations covering 15 districts of the country have revealed that the average background concentration of As in Bangladesh is much below 10 mg/kg soil. However, in some areas where soils receive Ascontaminated groundwater irrigation, the concentration has been found to be as high as 80 mg/kg soil. The soil As varies both spatially and vertically. The soil formation and the aquifer characters control the spatial variation, while the vertical distribution is controlled by the clay contents. The maximum As concentration in irrigation water was found to be 0.55 mg/L; irrigating a rice field with this water when the requirement is 1000 mm of water, it has been calculated that the As load will come to 5 kg As/ha/yr. Many crops receiving As contaminated water as irrigation have been found to accumulate As at levels that exceed the maximum allowable daily limit (MADL) of 0.2 mg per kg dry weight (dw). Some vegetables crops like Arum (Colocassia antiquorum), Kalmi (Ipomea aquatica), Amaranthus (Amaranthus spp.) etc. were found to be As accumulators. The transfer factor for As has been found to exceed the value of 0.1 in a number of plants indicating their affinity towards this element. In Arum, the concentration of As have been found to be as high as more than 150 mg/kg dw. Rice and wheat receiving As-contaminated irrigation water have been found to sequester the toxic metalloid into roots and stems. However, the quantities of rice consumed per person per day with the content of As in the grain may in many instances, surpass the MADL.

1

INTRODUCTION

Contamination of groundwater by arsenic (As) in the deltaic region, particularly in the Gangetic Alluvium of Bangladesh has become one of the world’s most important natural calamities. In many areas of the country, water containing more than 0.05 mg/L As, a limit value set for drinking water has been reported (DPHE-BGS 1999). Efforts are being directed towards ensuring safe drinking water either through mitigation technique or through finding alternative sources. Even if an As-safe drinking water supply is ensured the same groundwater will continue to be used for irrigation 95

purpose, leaving a risk of soil accumulation of this toxic element and eventual exposure to the food chain through plant uptake and animal consumption. Between 30 to 40% of net cultivable land is under irrigation and more than 60% of this irrigation is met from groundwater. Studies in the last few years have focused on ingestion of As through intake of groundwater containing As. The observation that As poisoning among the population is not consistent with the level of water intake has raised question on potential pathways of As ingestion. In a preliminary study from a village of Laksimpur, Huq et al. (2001) observed that certain vegetables growing on soil supposed to be affected by As contamination, accumulated substantial amount of As in them. This observation prompted the authors to undertake further study to see the loading of arsenic in soils from irrigation water and subsequently to different crops growing on those soils.

2

MATERIALS AND METHODS

For the study, information on As contamination in Bangladesh was obtained from secondary sources (DCH 1997, DPHE/BGS 1999). On the basis of the information, some areas were identified as control (where groundwater As contamination was not reported), less affected, moderately affected, severely affected (according to the incidence of As patients). Thus water, soil, and vegetables/crop samples were collected from 143 locations, covering 13 districts of Bangladesh (Huq & Naidu 2003). Water samples from hand tube-wells or irrigation pumps were collected. Soon after collection of water samples 2–3 drops of concentrated HCl was added to the vials containing water and transported to the laboratory for further analysis. Surface (0–15 cm) and subsurface (15–30 cm) soil samples were collected in replicate from locations covering the alluvial flood plains of Ganges, Teesta and Meghna-Brahmaputra rivers as well as from the Pleistocene terraces (Huq et al. 2003). Arsenic poisoning was reported from all these areas except the Pleistocene terraces, which was taken as control for comparison. In order to monitor the As load on soils from water, samples were collected from regions subjected to hand tube-well, shallow tube-well, and surface water irrigation. Random grid sampling was adopted and from each site the number of samples collected ranged from 25 to 40 per acre. After collection, samples were air dried, ground and screened to pass through 0.5 mm sieve and stored in plastic vials and set for further analysis. Replicate samples of consumable parts of the vegetables/crops commonly grown in the sampling area were collected. All plant samples were dusted free of adhering soil particles, washed with deionized water and 0.05 M HCl and then washed with deionized water 3 times to ensure dislodging of adhering dust particles. Samples were then dried in oven at 60 ⫾ 5°C for 48 h, ground, screened to pass through 0.2 mm sieve, stored in plastic vials, and kept for further analysis. The As in soil was extracted by digesting with aqua regia while As in plant samples was extracted with HNO3 digestion (Portman & Riley 1964). Arsenic in water, soil extract, and plant extract was estimated by HG-AAS technique and certified reference materials were used to ensure QA/QC.

3

RESULTS AND DISCUSSION

The As concentration in water used for irrigation was found to vary between 0.14 to 0.55 mg/L. So, for a Boro rice requiring 1000 mm of irrigation water per season, the load of As comes to between 1.36 to 5.5 kg/ha/yr. Similarly, with winter wheat requiring 150 mm of irrigation water per season, the load of As is estimated to be between 0.12 to 0.82 kg/ha/yr. The results on soil As indicated that the collected soils were not contaminated with As, and contained less than 10 mg/kg As. Expressed in terms of kg per hectare; the values did not exceed 20 kg/ha (Fig. 1). Moreover, the As concentration was higher in the surface layer than the sub-surface layer with a few exceptions in soils collected from Meghna Alluvium. On the other hand, soils from arsenic contaminated area show higher values ranging from around 2 to ⬎80 mg/kg. It needs to be mentioned 96

20

Soil As (kg/ha)

15 0-15 cm

15-30 cm

10

Figure 1.

Gangetic Alluvium soil

Megna Alluvium soil

Teesta Alluvium soil

0

Old Pleistocene Alluvium soil

5

Arsenic content in different soils of Bangladesh. Table 1. Arsenic content (mg/kg dw) in common plants from uncontaminated and contaminated areas of Bangladesh. Name of plants

Uncontaminated areas

Contaminated areas

Green papaya (Carica papaya) Arum (Colocassia antiquorum) Bean (Dolicos lablab) Indian spinach (Brasilia alba) Long bean (Vicia faba) Potato (Solanum tuberosum) Bitter gourd (Momordicum charantia) Aubergine (Solanum melongena) Chili (Capsicum spp)

0.46 0.39 0.092 0.15 0.30 0.62 1.56 0.23 0.41

2.22 153.2 1.16 1.00 2.83 2.43 2.12 2.3 1.52

here that soils belonging to the Gangetic Alluvium contained higher amount of As than the soils belonging to the Teesta Floodplain. Usually in soils contaminated through anthropogenic activity the arsenic contents may rise up to 50 mg/kg. In the present study, although the source of As is geogenic (Nickson 1998) yet, in some cases the values were equal to that of anthropogenic activity. This is an indication of As accumulation in the soil due to irrigation. Results on As contents of the analyzed vegetables/crops showed that some of the vegetables/crops accumulated As in the plant tissues (Table 1). It is also apparently clear that plants of the same type growing on uncontaminated soil had much less As content in their tissues. For rice and wheat, most of the As taken up by the plants were sequestered in the roots and stems, an insignificant amount was found to have accumulated in the grains (Fig. 2). But some leafy vegetables, particularly arum (Colocassia antiquorum) appeared to be an As accumulator. This was true for all the areas studied (Fig. 3), indicating that As from groundwater was entering into the food chain through soil to crops transfer. The levels of As in plants seldom exceed 1 mg/kg (Markert 1992). In the present study, quite a few plant samples had values much higher than this level. Farago & Mehra (1992) have considered that when the plant/soil ratios for any particular element are 0.1, then the plant can be considered as excluding the element from its tissues. In our case, many plants have shown this phenomenon while some like arum and a number of leafy vegetables had shown the reverse phenomenon indicating 97

As content (mg/kg dw)

80 Average Maximum Minimum

60

40

20

0 Grain

Figure 2.

Husk

Leaf

Stem

Root

Arsenic content in different parts of rice plant. 10

As content (mg/kg dw)

9

Plant of Gangtetic Alluvium soil

8

Plant of Meghna Alluvium soil

7

Plant of Teesta Alluvium soil Plant of Pleistocene terrace soil

6 5 4 3 2 1 0 um

Ar

al

inj

Br

us

nth

ra ma

urd

Go

ion

On

A

Figure 3.

a

h

nac

pay

Pa

n dia

spi

h

dis

Ra

In

Arsenic content in different crops collected from soils of various origin.

their affinity to As accumulation (Table 2). It was observed that arum (Colocassia antiquorum) showed very high accumulation of As when grown with As contaminated water. This plant grows in wet areas. As a result it is all the time taking up As from groundwater. Moreover, this plant is consumed very widely in the rural areas of Bangladesh. It is a good source of vitamin A, C and Fe. Every parts of the plant such as leaves, stems, rhizomes, and creepers are edible and are consumed. For this reason, this plant was thoroughly analyzed. The average As content in arum plants collected from different areas of Bangladesh is presented in Table 3. The maximum allowable daily level of As in foodstuff is taken as 0.22 mg per day (OEHHA 2003). On the basis of this level, calculations were made on the possibility of exceeding this MADL for the various plants analyzed. For example, a person who daily consumes 100 g of arum that contains 2.2 mg/kg of As would have a MADL from arum alone. However, when the concentration 98

Table 2.

Arsenic transfer factors in different plants of Bangladesh.

Name of crops

Arsenic transfer factor

Name of crops

Arsenic transfer factor

Mustard Patal Chicinga Ladies finger Coriander Kalmi Pui sak Cow-pea Pumpkin Lentil Radish Chili Carrot Papaya

0.02 0.08 0.11 0.11 0.12 0.12 0.12 0.13 0.14 0.17 0.18 0.2 0.23 0.26

Bean Karalla Brinjal Cabbage Amaranthus Tomato Garlic Turmaric Gourd Jhinga Pea Cauliflower Wheat Arum

0.27 0.32 0.35 0.44 0.55 0.55 0.57 0.68 0.69 0.69 0.83 0.84 1.46 2.64

Table 3.

Arsenic content in arum collected from different locations of Bangladesh. As content (mg/kg dw)

Location

Average

Maximum

Minimum

Brahmanbaria Chuadanga Comilla Dhaka Dinajpur Jessore Meherpur Munshigonj Narayangonj Pabna Rangpur

23.55 0.99 2.03 0.40 0.15 5.27 0.71 0.78 3.08 24.65 1.10

138.33 3.78 4.66 0.96 0.21 11.37 1.52 3.05 20.50 115.32 3.82

0.83 0.13 0.03 0.05 0.09 0.92 0.15 0.18 0.02 1.05 0.14

is as high as 22 mg/kg, only 10 g would give the MADL. Similarly, 440 g of rice with 0.5 mg/kg would also represent the MADL. Such inputs are comparable to drinking 4.4 L of water with 0.05 mg/L. On the basis of the As content in rice, the amount of average daily consumption, the extent of As contamination in the area, the incidence of arsenicosis patients and the number of population at risk to exposure of arsenic ingestion, the dietary load estimation, i.e., the possibility of the per cent of population risking the exposure to excess of MADL, has been calculated for Jessore (representing Gangetic Alluvium) and Rangpur (representing Teesta Alluvium) areas. In Jessore area, 32% of the people are above the MADL, while in Rangpur area the value is only 2% (Huq et al. 2001). This again substantiates the fact that the groundwater in the Gangetic alluvial plain is more contaminated than the other parts of Bangladesh. Cooked food from the households of arsenic affected people were collected and analyzed. Arsenic contaminated water has been used to cook these foods. The foods contained various amount of As in them (Table 4). It is interesting to note that As could not be detected in cooked lentil soup, locally called “Dal” and in eggs. The cooked rice contained different amounts of As; the differences could be due either to the variety of rice and also due to variation in the As content in the cooking water. It can be seen that in many of the cooked foods the values are well above the 99

Table 4. Arsenic content (mg/kg) in some cooked food collected from various Arsenic contaminated areas. Sample

As

Sample

As

Rice Rice Rice Rice Vegetable curry Vegetable curry Spinach Spinach Spinach Spinach

0.35 0.36 0.11 0.13 0.81 0.95 0.13 0.12 0.34 0.33

Egg Egg Eggplant Eggplant Lentil soup Lentil soup Pumpkin Pumpkin Fish curry Fish curry

⬍b.d.l. ⬍b.d.l. 0.66 1.45 ⬍b.d.l. ⬍b.d.l. 0.27 0.25 0.34 0.39

⬍b.d.l. ⫽ below detection level.

MADL of 0.22 mg/kg. This is a matter of concern and it is an indication that As ingestion in human beings is affected not only through water but also through foodstuffs. The food habit and the nutritional status of a person thus could be related to the manifestation of arsenicosis. This, however, is also related to the biomethylation activity of the individual (Alauddin et al. 2002). The above information indicates that there are other pathways of As ingestion in human body besides drinking water, and that is through food chain. Crops receiving As contaminated irrigation water take up this toxic element and accumulate it in different degrees depending on the species and variety as well as on the type of soils on which these plants are growing. However, the portion of this As that goes directly to the different metabolic pathways and causes the problem of arsenicosis needs to be assessed. The bioavailability of this arsenic in the different food materials also needs to be assessed. In a preliminary study by the authors with swine feeding trials, it has been observed that 27% of the total amount of As in silverbeet and 82% of As in rice were bioavailable.

ACKNOWLEDGEMENTS The present work is a part of a joint research financed by Australian Center for International Agricultural Research (ACIAR) and the Ministry of Education, Government of Bangladesh.

REFERENCES Alauddin, M., Chowdhury, D., Bhattacharya, M., Bibi, H., Begum, S., Islam, M.S. & Rabbani, G. 2002. Speciation of arsenic metabolic intermediates in human urine by chromatography and flow injection hydride generation atomic absorption spectrometry. Paper presented at the 4th International Conference on “Arsenic Contamination of Ground Water in Bangladesh: Cause, Effect &Remedy” held at Dhaka, Bangladesh during 12–13 January 2002. DCH (Dhaka Community Hospital) 1997. Arsenic pollution in groundwater in Bangladesh, Dhaka, Bangladesh. DPHE/BGS 1999. Groundwater studies for arsenic contamination in Bangladesh, Phase 1: Rapid Investigation Phase, Final Report in 4 volumes prepared for the Government of Bangladesh and the Department for International Development (UK). Farago, M.E. & Mehra, A. 1992. Uptake of elements by the copper tolerant plant Armeria maritima. Metal compounds in environment and life 4, (Interrelation between Chemistry and Biology), Science and Technology Letters, Northwood. Huq, I., Smith, E., Correll, R., Smith, L., Smith, J., Ahmed, M., Roy, S., Barnes, M. & Naidu, R. 2001. Arsenic Transfer in Water-Soil-Crop Environments in Bangladesh I: assessing Potential Exposure Pathways in

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Bangladesh. Book of abstracts, “Arsenic in the Asia-Pacific Region Workshop Adelaide 2001”, held during 20–23 Nov 2001, Adelaide, South Australia. Huq, S.M.I., Jahan Ara, Q.A., Islam, K., Zaher, A. & Naidu, R. 2001. The possible contamination from arsenic through the food chain. In Groundwater Arsenic Contamination in the Bengal Delta Plains of Bangladesh (Proceedings of the KTH-Dhaka University Seminar, University of Dhaka, Bangladesh), Eds. G Jacks, P Bhattacharya, and AA Khan KTH Special Publication, TRITA-AMI Report 3084, ISSN 14001306, ISRN KTH/AMI/REPORT 3084-SE, ISBN 91-7283-076-X, ©2001, KTH, pp. 91–96. Huq, S.M.I. & Naidu, R. 2003. Arsenic in groundwater of Bangladesh: Contamination in the food chain. In M.Feroze Ahmed (ed.), Arsenic Contamination: Bangladesh Perspective.ITN-Bangladesh, BUET, Dhaka, Bangladesh, ISBN 984-32-0350-X, pp. 203–226. Huq, S.M.I., Rahman, A., Sultana, N. & Naidu, R. 2003. Extent and severity of arsenic contamination in soils of Bangladesh. In M.Feroze Ahmed, M.Ashraf Ali and Zafar Adeel (eds.), Fate of Arsenic in the Environment. BUET, Dhaka , The United Nations University, Tokyo, ISBN 984-32-0507-3, pp. 69–84. Merkert, B. 1992. Multi-element analysis in plant materials – Analytical tools and biological questions. In DC Adriano (ed), Biogeochemistry of Trace Metals, Lewis Publishers, Boca Raton. Nickson, R., Mcarthur, J.M., Burgess, W, Ahmed, K.M., Ravensroft, P. & Rahman, M. 1998. Arsenic poisoning of Bangladesh groundwater. Nature 395: 338. OEHHA 2003. Maximum allowable dose level (MADL) for reproductive toxicity for arsenic (inorganic oxides) for oral exposure. Proposition 65, Office of Environmental Health Hazard Assessment, Reproductive and Cancer Hazard Assessment Section, May, 2003. 7p. Portman, J.E. & Riley, J.P. 1964. Determination of arsenic in seawater, marine plants and silicate and carbonate sediments. Anal. Chem. Acta. 31: 509–519.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Arsenic contamination in groundwater in Nepal: a new perspective and more health threat in South Asia S.R. Kanel, H. Choi, K.W. Kim, and S.H. Moon Gwangju Institute of Science and Technology, Puk-gu, Gwangju, Republic of Korea

ABSTRACT: A study on arsenic (As) contamination of groundwater in the Gaur Municipality, Rautahat district of Nepal was undertaken during October, 2003. This study investigates the occurrences and distribution of As and its mechanism of release in groundwater in the affected areas. Our observations reveal that the groundwater was alkaline with high concentration of bicarbonates (562.8 mg/L), rich in iron, manganese, and silica Analyses of As in groundwater samples from 50 (both private as well as public) shallow tubewells revealed concentrations in the range between 1–62 ␮g/L. Among analyzed samples, 1 exceeded 50 ␮g/L concentration and 18 were between 10–50 ␮g/L concentrations and rest of samples (31) were below 10 ␮g/L concentration. However, the WHO guideline of maximum concentration limit of As in drinking water is 10 ␮g/L. The release of As from the sediments is related to: (i) the reduction of iron oxides and hydroxides, found to be the main mechanism of mobilization of As in the groundwater; (ii) desorption of As by phosphate and silica; and (iii) anion exchange with OH⫺. Arsenic concentration found in tubewells indicates that people are consuming As contaminated water (without any pretreatment) at serious risk of As poisoning. However, there is no counter treatment of As diseases; As remediation is the only one option to save the lives of people.

1

INTRODUCTION

Arsenic (As) is a metalloid element, within group V(B) of periodic table, present in natural water in a variety of forms (organic and inorganic), oxidation states according to pH and redox conditions (Ferguson & Gavis 1972). Since As is carcinogenic and extremely toxic, it may cause neurological damage at aqueous concentrations as low as 10 ␮g/L (Chunming & Robort 2001). The guideline value for As in drinking water recommended by World Health Organization (WHO) is 10 ␮g/L (WHO 2001). Keeping this in view efforts have been made to analyze As in groundwater in Nepal. Sandwiched between India and China, Nepal is a small mountainous country. The Terai region (Nepal-Gangetic Plain (NGP), forms the northern extension of the Indo-Gangetic Plain consists of 20 districts in southern part bordering India, about 33 km from Bangladesh, contains about 40 km width and 885 km length, and its geology is similar to the Bengal Delta Plain (BDP) in Bangladesh and West Bengal, India (Tandukar et al. 2001). Very recently, As has been detected in groundwater from the alluvial aquifers of southern Nepal in Terai region (Tandukar 2000, Tandukar et al. 2001, 2004 this volume, Bhattacharya et al. 2003). The objective of this study is to locate the areas affected by natural As occurrences in groundwater of Nepal, and to understand the release mechanisms of As within the study area situated in Rautahat District, Gaur Municipality in southern part of Nepal bordering India. 2

MATERIALS AND METHODS

Fifty goundwater samples were collected randomly from private and public tube wells in the Gaur municipality of the Rautahat District, Nepal in October 2003 (Fig. 1). A typical hand pumped 103

tubewell is shown in Figure 2 where a people collect groundwater for drinking purposes. The generally crystal clear water samples were collected in 50 mL polypropylene centrifuge tubes, which were sealed airtight in spot (Berg et al. 2001). Water samples were filtered on site using 0.45 mm membrane filters for anion analyses and samples for cation analyses were acidified by concentrated nitric acid. The well positions were located with the help of topographical map of Rautahat district. The depth of well, pH, oxidation-reduction potential (ORP), conductivity and total dissolved solids (TDS) were measured immediately at the site. Water samples were analyzed for total As and cations (total iron, sodium, potassium, manganese and calcium) by atomic absorption

Figure 1.

Location map of the area of study in Rautahat district of Nepal.

Figure 2.

Collection of groundwater from tubewell in Gaur Municipality, Rautahat District, Nepal.

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spectroscopy (AAS) using a Solaar instrument (Britain). For As measurements, an on-line hydride generation device was coupled to the AAS (HG-AAS). Anions (bicarbonate, chloride, sulfate and nitrate) were analyzed using standard methods (APHA, AWWA, WPCF, 2000). The ORP, pH, conductivity, TDS were measured by ultra meter (Ultra meter TM, model 6p). 3 3.1

RESULTS AND DISCUSSION Field Parameters

Groundwater temperatures ranged between 25 to 29.1°C. The temperature of the most of the As contaminated groundwater was found at 25.5 to 27.5°C. The depth of the tubewells ranges from 4 to 107 m. Most of the As exists in 20 to 100 ft depth. The pH values of samples ranged from 7.29 to 7.80. The TDS of groundwater varied from 263 to 1987 mg/L. The ORP values ranges form 81 to 142 mV. The conductivity of groundwater varied from 283 to 2457 ␮S/cm. 3.2

Major ion chemistry of groundwater and its relation to arsenic

Major ion concentrations in the groundwater from Gaur municipality are shown in Figures 3 and 4. The groundwater pH is neutral to slightly alkaline. Even though the ORP value is positive (81–142 mV), the condition of groundwater is fluctuating and remains to be in reducing condition. The 2⫺ ⫺ distribution of major anions such as NO⫺ 3 (0–45.9 mg/L), SO4 (0–17 mg/L), Cl (0–334 mg/L), 70

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Cation concentration in groundwater and its variability with the distribution of As.

HCO⫺ 3 (101–563 mg/L) show significant variations. The concentration of major cations such as Na (10–134 mg/L), K (1–271 mg/l), Ca (50–214 mg/L indicated considerable variations. Similarly, the concentrations of total Fe (0.0–8.2 mg/L) and Mn (0–3.05 mg/L) were variable in the analyzed groundwaters. The groundwater is predominantly Ca-HCO3 type with high concentration of bicarbonate ions. The groundwater of West Bengal, India was found to be Ca-HCO3 type (Stüben et al. 2003), whereas in Bangladesh it was also found to be Ca-HCO3 or Ca-Mg-HCO3 type (Ahmed et al. 2004). Nepal shares similar geological characteristics of the aquifers with West Bengal, India and Bangladesh. Concentration of total As in Gaur Municipality varied from 1 to 62 ␮g/L. The concentrations of As at different tube wells were plotted with anions and cations (Figs 3 and 4). In Rautahat district, As concentration of 36% of the tubewell samples exceeded WHO guideline value of 10 ␮g/L and 2% samples exceeded the Nepalese Interim standard of 50 ␮g/L. The average concentration of phosphate, ammonia and silica was found to be 0.2, 0.11 and 39.36 mg/L in the samples of which As concentrations was very high. 3.3 Mechanism to release arsenic in groundwater It is evident that the distribution of As in the groundwater is governed by a complex interaction of different factors, among which microbially mediated redox process, adsorption/desorption, precipitation and ligand exchange are the most important. which reduction of oxy-hydroxides may be the main factor. Based on the above results, and taking into account the redox chemistry of As, reduction of oxy-hydroxides may be suggested to explain the distribution of As concentrations in the groundwater in Nepal of the study area. Under higher redox levels (⫹200 to ⫹500 mV), As will be adsorbed predominantly as As(V), which is more strongly adsorbed on ferrihydrite as compared to As (III) (Pierce & Moore 1982). Additionally, a Fe (III)–As (III) complex is more soluble than a Fe (III)–As(V) complex (Gulens et al. 1973). At gradually decreasing redox potential (e.g., due to bacterial activity, lack of O2) As (V) will be reduced to As (III) (Masscheleyn et al. 1994), which would imply the mobilization of a part of the As, because of the higher mobility and weaker bond of As (III) on ferrihydrite. If the redox potential decreases below ⬍100 mV, ferrihydrite will dissolve, releasing all attached As (Masscheleyn et al. 1994, Deutsch 1997). Even though the ORP value of groundwater was found to be 81–142 mV, low nitrate (45.9 ␮g/L), relatively high Fe (10.4 mg/L) and Mn (3.1 mg/L) and high concentration of As suggest that the groundwater in the well is reducing (Smedley et al. 2003). In this condition the dissolution of oxy hydroxides produce Fe2⫹ and HCO⫺ 3 ions according to following reaction (Nickson et al. 2000):

The present data suggest poor correlation between Fe2⫹ and HCO⫺ 3 (data plots not shown) that may be explained by the possible formation of Fe2CO3 (Sracek et al. 1998, Welch & Lico 1998, Ahmed et al. 2004). As the concentration of total As increases, there is constant release of manganese. There may be constant release of As through manganese but the increase of As concentration may be due to iron oxyhydroxides. (Bhattacharya et al. 2003) has also reported mobilization of As as a desorption of As-oxyanions adsorbed into Fe and Mn oxides as well as reductive dissolution of these surface reactive phases from the sediments in Nawalparashi District, Nepal (Bhattacharya et al. 2003). Stüben et al. (2003) has also reported mobilization of As. In West Bengal, India by dissolution of iron oxyhydroxides (Fe(III) to Fe(II)) and Mn phases due to the decreasing reducing contion of groundwater (Stüben et al. 2003). 4

CONCLUSIONS

Source of As in the groundwater of Nepal is natural. About 36% of shallow tubewells in Rautahat district, Nepal was found to exceed the maximum permissible limit for safe drinking by World 107

Health Organization (10 ␮g/L, WHO 2001). Since the bicarbonate concentration are high and on the contrary sulfate concentration was found as very low, it implies that the mobilization of As in the groundwater is not governed by oxidation of arsenopyrite. However the groundwater was found to have high bicarbonates and iron concentration, which implies that the release of As in the groundwater was due to reduction of iron oxy-hydroxides. Moreover, As was also released due to dissolution of manganese oxyhyroxides, which is confirmed by the presence of manganese in all the tubewells.

REFERENCES APHA, AWWA WPCF. 2000. Standard Methods for the Examination of Water and Wastewater, 20th Edition. Ahmed, K.M. Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Hossain, M.A., Bhuyian, M. Imam, B., Khan, A.A. & Sracek, O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh: an overview. Appl. Geochem. 19(2): 163–260. Bhattacharya, P., Tandukar, N., Neku, A., Valero, A.A., Mukherjee, A.B. & Jacks, G. 2003. Geogenic arsenic in groundwaters from Terai Alluvial Plain of Nepal. J. Phys. IV France 107: 173–176. Chunming, S. & Robort, W.P. 2001. Arsenate and arsenite removal by zerovalent iron: kinetics, redox transformation, and implications for in-situ groundwater remediation. Env. Sci. Tech. 35: 1487–1492. Deutsch, W.J. 1997. Groundwater Geochemistry: Fundamentals and Applications to Contamination. Lewis Publishers,Boca Raton, New York. Ferguson, J.F. & Gavis, J. 1972. Review of the arsenic cycle in natural waters. Water Res. 6: 1259–1274. Gulens, J., Champ, D.R. & Jackson, R.E. 1973. Influence on redox environments on the mobility of arsenic in groundwater (Chapter 4). In: Rubia, A.J. (ed.) Chemistry of Water Supply Treatment and Distribution. Ann Arbor Science Publishers, Ann Arbor, MI. Masscheleyn, P.H. & Patrick, W.H. Jr., 1994. Selenium, arsenic and chromium redox chemistry in wetland soils and sediments. In: D.C. Adriano, Z.-S. Chen, S.-S. Yang (Eds.), Biogeochemistry of Trace Elements. Environ .Geochem .Health 16 (Special issue), 615–625. Berg, M., Tran, H.C., Nguyen, T.C.H., Pham, H.V., Schertenleib, R. & Giger, W. 2001. Arsenic contamination of groundwater and drinking water in Vietnam: A human health threat. Env. Sci. Tech. 35(13): 2621–2626. Nickson, R.T., McArthur, J.M., Ravenscroft, P., Burgess, W.G. & Ahmed, K.M. 2000. Mechanism of arsenic release to groundwater, Bangladesh and West Bengal. Appl. Geochem. 15: 403–415. Pierce, M.L., Moore, C.B., 1982. Adsorption of arsenite and arsenate on amorphous iron hydroxide. Water. Res., 16: 1247–1253. Smedley, P.L., Zhang, M., Zhang, G & Luo, Z., 2003. Mobilization of arsenic and other trace elements in fluviolacustrine aquifers of the Huhhot basin, Inner Mongolia. Applied geochemistry, 18: 1453–1477. Sracek, A., Bhattacharaya, P., Jacks, G., Chatterjee, D., Larson, M. & Leiss, A., 1998. Groundwater As in the Bengal Delta Plains: a sedimentary geochemical overview. In: Int. Seminar Applied Hydrochemistry, Annamalai Univ., Tamil Nadu, India. 18–20. Stüben, D., Berner, Z., Chandrasekharam, D. & Karmakar, J. 2003. Arsenic enrichment in groundwater of West Bengal, India: geochemical evidence for mobilization of As under reducing conditions. Appl. Geochem. 18: 1417–1434. Tandukar, N. 2000. Arsenic contamination in ground water in Rautahat District of Nepal. An assessment and treatment, MSc Thesis Institute of Engineering, Tribhuvan University, Lalitpur, Nepal (unpublished). Tandukar, N., Bhattacharya, P. & Mukherjee, A.B. 2001. Preliminary assessment of arsenic contamination in groundwater in Nepal. Book of Abstracts, Arsenic in the Asia-Pacific Region Workshop, CSIRO, Adelaide, Australia:103–105. Tandukar, N., Bhattacharya, P., Jacks, G. & Valero, A.A. 2004. Naturally occurring arsenic in groundwater of Terai region in Nepal and mitigation options. In: Bundschuh, J., Bhattacharya, P. & Chandrashekharam (eds.) Natural Arsenic in Groundwater: Occurrences, Remediation and Management, Taylor and Francis Publications (This Volume). Welch, A.H. & Lico, M.S. 1998. Factors controlling As and U in shallow groundwater, southern Carson Desert, Nevada. Appl. Geochem. 13: 521–539. WHO 2001. Arsenic in drinking water. Fact sheet 210: URL: http://www.who.int/mediacentre/factsheets/ fs210/en/print.html (Accessed on March 9, 2004).

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Estimating previous exposure to arsenic for populations living in parts of Hungary, Romania and Slovakia R.L. Hough, G.S. Leonardi & T. Fletcher London School of Hygiene & Tropical Medicine, University of London, London, UK

ABSTRACT: Arsenic Health Risk Assessment and Molecular Epidemiology (ASHRAM) is a large EU-funded case-control study looking at risk of cancer (bladder, kidney and skin) due to arsenic in drinking water. The ASHRAM project is specifically concerned with populations residing in parts of Hungary, Romania and Slovakia. In these districts, many incidences of exposure to arsenic (As) have occurred in the recent past. However, some populations are continuing to experience exposure to relatively low-level arsenic concentrations in their drinking water. In order to calculate risk of cancer, it is necessary to estimate cumulative arsenic dose by reconstructing individual exposure histories due to water consumed both at home and in the workplace. In this study, data on water consumption was obtained using questionnaire tools. Concentrations of arsenic in water supplies were measured using Atomic Absorption Spectrophotometry (AAS). Measurements of total inorganic arsenic in the urine of the participants were made as a biomarker for current exposure. Preliminary results indicate that on average people are drinking 1.76 L/d from tap sources. Initial analysis suggests that the volume of water consumed by individuals has changed little with time, while the majority of people drink between 10 and 25% of their water at work. These estimates seem plausible and similar to other estimates.

1 1.1

INTRODUCTION Background

One of the main difficulties in epidemiological studies which examine the relationship between drinking water contaminants, and disease outcomes, is the accurate assessment of individuals’ consumption of tap water (Shimokura et al. 1998). A recent workshop identified improving methods for measuring water consumption patterns as a primary research need for epidemiological studies (Arbuckle et al. 2002). Although this workshop was primarily concerned with the health risk associated with disinfection by-products in domestic water, their conclusions are relevant to the study of other drinking water contaminants. Humans may ingest water either directly or indirectly. Direct water intake refers to plain water which is ingested as a beverage (USEPA 2000). Indirect water intake is the ingestion of water added to food as part of preparation, excluding water that is innate to foods (USEPA 2000). Total fluid intake is the ingestion of all fluids including water innate to foods and drinks such as juices and sodas (USEPA 1997). Total tap water intake refers to the ingestion of tap water for both drinking and food preparation purposes (USEPA 1997). In this study, we aim to estimate the total tap water intake at an individual level. A number of studies have attempted to assess total tap water intake among a variety of different populations (e.g. Bates et al. 2004, Steinmaus et al. 2003, Williams et al. 2001, USEPA 2000, Swan & Waller 1998, USDA 1995, Roseberry & Burmaster 1992, Ershow et al. 1991, Ershow & Cantor 1989). In general, large national surveys tend to report relatively lower estimates of water intake compared with estimates from smaller studies. For example, small studies looking at specific populations, 109

e.g. for epidemiological research, tend to report relatively higher estimates of water intake. Table 1 provides a summary of selected water intake studies (the USEPA’s Exposure Factors Handbook provides a more comprehensive summary of water intake studies dating back to 1968, USEPA 1997). However it is difficult to make comparisons between studies of water intake because they vary widely in methodology, sample size, population and the type of water intake estimates obtained/derived. 1.2

ASHRAM project

Arsenic (As) is a widespread element and is present at elevated concentrations in natural underground water in some areas of Europe. It is known to cause several different types of cancer, but the magnitude of risk associated with the relatively low-level exposures that occur in Europe is still uncertain. The ASHRAM (Arsenic Health Risk Assessment and Molecular Epidemiology) study is a case-control study which has been funded under Key Action 4: Environment and Health in the European Union Quality of Life section of the Fifth Framework programme to investigate the nature of arsenic-related cancer risk (bladder, kidney and skin) at concentrations relevant to millions of citizens in the European Union and ‘Accession Countries’. The ASHRAM project is specifically concerned with populations residing in parts of Hungary, Slovakia and Romania (Fig. 1). In these districts, many incidences of exposure to arsenic have Table 1.

Summary of water intake estimates from recent studies.

Study USEPA (2000) Foundation for Water Research (1996) USDA (1995) Ershow and Cantor (1989) Ershow et al. (1991) Cantor et al. (1987) Canadian Ministry of National Health & Welfare (1981) Bates et al. (2004)

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Figure 1. Map of Central Europe showing the ASHRAM study areas (highlighted) in parts of Hungary, Romania and Slovakia.

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occurred in the recent past. However, some populations are continuing to experience exposure to relatively low-level arsenic concentrations in their drinking water. For many people living within the study areas, the main sources of drinking water include wells or standpipes located in the street or village square. The majority of these sources have provided drinking water over many years, and often serve whole communities. Exposure may occur where wells have been drilled into aquifers that contain arsenic, or where water had elevated concentrations of arsenic in the past.

2

METHODOLOGY

For 1200 cancer cases (300 bladder, 300 kidney and 600 skin cancer) and 600 controls, estimates of cumulative arsenic dose will be constructed and cancer risks will be estimated in relation to arsenic intake (adjusting for potential confounding by diet, tobacco and occupation). The impact on risk of individual differences in arsenic metabolism, DNA repair and diet will be assessed. For sub groups of the population, more detailed analysis of inter-individual differences in speciation, the effect of other nutrients and minerals, and specific polymorphisms will inform models of metabolism. Risk management will be explored for those with the highest continuing exposure. In order to calculate risk of cancer, it is necessary to estimate cumulative arsenic dose by reconstructing individual exposure histories. This requires two elements: (i) a measurement/estimate of arsenic in the water that is/was drunk, (ii) an estimate of the amount of water consumed from each source, by each individual. 2.1

Assessing exposure to arsenic from drinking water

Information regarding current exposure to arsenic was ascertained using two questionnaire tools: a Main Questionnaire which contained sections on health, lifestyle, residential and occupational history; and a Food Frequency Questionnaire. The Main Questionnaire provides information on a person’s current drinking water supplies (both residential and occupational) and the locations of these supplies. It also provides information on the amount of water consumed by each individual while at work. This information was used to locate the current water supplies used by each participant. The concentration of inorganic arsenic in water samples, [Asw] (␮g/L), was determined in each water supply using Hydride Generation – Atomic Absorption Spectrophotometry (HGAAS). The Main Questionnaire also provides details of each participant’s previous residencies and work places. This information may be used to locate past water sources in order to sample these. Each participant in the study also provided a urine sample as a biomarker for current exposure to arsenic. Total inorganic arsenic in urine, [Asu] (␮g/L), was determined using High Performance Liquid Chromatography – Hydride Generation – Inductively Coupled-Mass Spectrometry (HPLC-HG-ICPMS). This method was also used to determine the relative concentrations of metabolites of arsenic present in the urine. This information will be used to inform models of arsenic metabolism and genetic differences in arsenic susceptibility. The Food Frequency Questionnaire provides information on the amount of tap-water consumed at home by each individual in the study, Iw (L/d). This included not only the consumption of plain tap-water, but also consumption of food and drink items which contained tap water such as tea, coffee and soup. Past exposure will be determined using the information provided on past residencies and workplaces from the Main Questionnaire. An attempt will be made to locate and sample all significant past water sources for each individual. Where this is not possible, the use of proxy measures such as geology may be used. In order to account for political changes in the study countries, the Food Frequency Questionnaire is divided into two main sections, the first dealing with a typical year before 1989 (when the study countries underwent significant political changes), and a typical year after 1989 (but before the participant developed the symptoms which have caused them to be referred to the hospital). The information provided by the Food Frequency Questionnaire may then be used in conjunction with 111

the measured concentrations of arsenic in individual’s water supplies, to estimate both current and cumulative exposure. 2.2

Estimating water consumption

The volume of water consumed by each individual while at a particular residence or workplace was determined with information from both questionnaires. For residential water consumption, pre- and post 1989 water consumption were distinguished. In the home, water consumption was reported as plain tap water, cold drinks made with tap water (e.g. syrup drinks), tea, coffee and soups. Tap water is also added to foods as part of certain recipes. In the workplace, water consumption was reported as the intake of plain tap water, tea, coffee and soup. The Food Frequency Questionnaire reports how many ‘glasses’, ‘cups’, or ‘bowls’ each individual consumed per week or per day. In order to convert this into a volume, it was necessary to make assumptions about the volume of these particular vessels. The volumes chosen are presented in Table 2 and represent typical volumes for these vessels. The same volumes were used to estimate the volume of water consumed in the workplace using information from the Main Questionnaire. During interviews, the interviewer shows the participant an example glass of 0.15 L and asks the participant to give their estimate of intake assuming this glass size, i.e. if they usually drink from a glass twice as big, the participant should report their estimate as roughly twice the number of glasses consumed. However, this does not prevent people under- or overestimating intake due to not taking the vessel size into account. A validation exercise will be carried out on a subset of the study population. This will look at variation in drinking vessel size in relation to reported consumption. It is aimed to incorporate this variation into final estimates of water consumption. 2.3

Determination of ‘exposure periods’

Data derived from the questionnaires may be used to divide each individual’s exposure histories into different ‘exposure periods’ relating to the different sources of water used throughout a participant’s lifetime. Exposure periods for water principally relate to changes in residence and changes in job. The measured concentration of arsenic in water collected from all water sources which relate to a single exposure period may be used to determine exposure, expressed as an ‘Average Daily Dose’, ADD (␮g/d), for a single exposure period: (1) The Cumulative Dose, CD (␮g) for each participant may then be estimated as simply the summation of all exposure periods. Other indices of exposure which are of interest to epidemiological analysis are the ‘Average Dose Rate’, ADR (␮/d). This is calculated as an individual’s cumulative dose divided by their entire exposure duration, T: (2) Cumulative water intake, CI (L) may be calculated for each participant as the summation of intake across all work periods. The average drinking water consumption may then be calculated by Table 2. Drinking vessels and their subsequent assumed volumes used to convert information from the questionnaires into volumes of water consumed by study participants. Vessel

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dividing CI by the total duration of all work periods. It is then possible to divide the ADR for each individual by the average drinking water consumption. Where ‘average drinking water consumption’ may be for an individual or for a population.

3

PRELIMINARY RESULTS

Data are currently being collected; however preliminary data from the first 166 questionnaires are presented here to illustrate estimates of the consumption of water derived from major sources of interest (private well, public water supply, workplace supply) by each individual. At present, results relate only to each individual’s current residential and occupational water sources. 3.1

Concentration of arsenic in current water supply

Figure 2 a, b and c shows the preliminary results for concentrations of arsenic measured in participant’s current water supplies for the three study countries. The greatest current exposures are seen

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Figure 2. Histograms showing the concentration of arsenic (␮g/L) in current water supplies in (a) Hungary, (b) Romania and (c) Slovakia; and concentrations of arsenic (␮g/L) in urine samples from participants residing in (d) Hungary, (e) Romania and (f) Slovakia. X-axis values represent the maximum value in each category.

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Figure 4. Histogram showing the average daily intake across all exposure periods, as reported by questionnaire. X-axis values represent the maximum value in each category.

in Hungary with the majority of samples measured so far between 32 and 64 ␮g/L. Current exposures in Romania and Slovakia are generally much lower (between 0.5 and 8.0 ␮g/L). Figure 2 d, e and f show the preliminary results for concentrations of arsenic in participant’s urine. Although these data do not necessarily correspond with the data shown in Figure 3 a, b or c, it can be seen that in Hungary where the current exposure is greatest, the concentrations of arsenic in urine are also greater. 3.2 Water consumption Figure 3 shows the total number of water sources per individual, broken down into residential and occupational water sources, as derived from the Main Questionnaire. In the case of occupational water sources, the questionnaire reports the number of occupational periods. It has been assumed that each occupational period relates to one water source. Figure 4 shows the daily intake for each participant averaged across all residencies and jobs. From these data, the mean daily intake of tap water by participants in the ASHRAM study is 1.76 L/d, with a 90th percentile of 2.78 L/d. Figure 5 shows the consumption of tap water in the workplace as a percentage of total tap water consumption. On average, participants drink 13.5% of their tap water in the workplace, with a 90th percentile of 28.4%. These data will be used to estimate cumulative exposure to arsenic in drinking water from both residential and occupational sources. 114

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Figure 5. Histogram showing the consumption of tap water at work as a percentage of total tap water consumption. X-axis values represent the maximum value in each category.

4 4.1

DISCUSSION Concentration of arsenic in current water supply

Preliminary investigations into the three study regions suggested that the greatest exposures would probably be found in Romania, while Slovakia and Hungary would have much lower exposures (although Hungary would have greater historical exposure). However, the concentrations of arsenic measured in current water supplies are, in general, a lot lower than expected. The majority of exposures are below current EU legislation of 10 ␮g/L. The highest levels of exposure appear to be in Hungary, with the lowest in Slovakia. Measurements of total arsenic in the urine of participants seem to be related to the concentration of arsenic in current water supplies. However, at this stage in the investigation, there are too few corresponding measurements to conclude this with certainty. 4.2 Water consumption The results suggest that average daily tap water consumption is 1.76 L/d. This is similar to estimates made by other studies of drinking water intake (Table 1). However, the 90th percentile from our study (2.78 L/d) is slightly greater than those reported in Table 1. This may be due to the fact that this study is only in the preliminary stages, the assumptions made in order to estimate water intake, or due to over reporting by some study participants. A validation exercise is planned to evaluate the ability of the questionnaires to report water intake. However, it may not be valid to make comparisons between this study and the studies reported in Table 1 due to methodological and population differences. A number of assumptions were made in order to estimate tap water intake in litres. The volume of water consumed by each individual was derived from assumptions relating to the volume of ‘typical’ drinking vessels (Table 2.). Studies conducted in other countries have tended to use larger drinking vessels as a ‘typical’. For example, Robertson et al. (2000) used a glass size of 0.25 L for a study looking at drinking water intake in Melbourne, Australia. Ryan et al. (2000) used a nominal serving size of 0.296 L per glass for a study conducted in Maryland, USA. The size of ‘typical’ glass used in this study was chosen after consultation; however variation in vessel size will also be assessed as part of the validation exercise. During interviews, participants are shown a 0.15 L glass in order for them to estimate consumption using this size of glass as the ‘units’. Therefore it is not known whether choosing a different glass size would have affected the final consumption estimates. 115

In order to estimate each individual’s exposure to arsenic, a number of assumptions will have to be made. Firstly it will be assumed that the concentration of arsenic measured in a water source has not changed over time, and may be used to estimate past exposures. Previous studies looking at temporal variability have produced opposing results. Few studies of the long-term variability of arsenic concentrations at a single well location have been reported. Karagas et al. (2001) analysed two tap water samples taken 3 to 5 years apart from each of 99 private wells in New Hampshire, USA. Little temporal variation in arsenic concentrations was found. Conversely, Focazio et al. (2000) analysed the variability in arsenic concentrations over time in 355 wells in USA (time period unspecified). The results suggested that concentrations of arsenic were temporally stable in 116 wells with arsenic concentrations in water ⬍1 mg/L. However, arsenic levels in wells were deemed to be temporally unstable at concentrations which would be deemed significant for exposure assessment. Hinkle & Polette (1999) found that al-though concentrations of arsenic in well water remained stable in the majority of wells, variation was as much as 50% from the mean in other wells. Neither study attempted to determine the cause of the variability. Another issue with estimating exposure histories is the fact that some individuals have lived for periods of time outside the study area. It will be assumed that the concentration of arsenic in drinking water from outside the study area is related to geology. It was also assumed that the water consumption reported in the Food Frequency Questionnaire was the same for all residencies, with the only distinction being pre- or post-1989. This assumption may be valid as a previous study (Levallois et al. 1998) found that total volume of water consumed in parts of Canada has changed little over time. However the relative importance of different beverages such as tea, coffee and bottled water has changed. It was also assumed that there was no change in water source at individual workplaces. The initial results suggest that the majority of people drink between 10 and 25% of their water at work, with a mean of 13.5%. These estimates seem plausible, given that similar results have been reported from Canada (Levallois et al. 1998) who reported that 25% of water intake occurred away from home, and in the United States Environment Protection Agency, USEPA, Exposure Factors Handbook (USEPA 1997) which suggests that 30% of water intake occurs away from home. These estimates may be larger than ours, but these are based on all water consumption away from the home (including non-tap water sources), and therefore do not simply focus on the workplace and total tap water intake.

5

CONCLUSIONS

At this preliminary stage, the estimates of tap water intake provided by the questionnaires used in the ASHRAM study seem reasonable and are similar to previous estimates. It was also assumed that the water consumption reported in the Food Frequency Questionnaire was the same for all residencies, with the only distinction being pre- or post-1989. This assumption may be valid as a previous study (Levallois et al. 1998) found that total volume of water consumed in parts of Canada has changed little over time. However the relative importance of different beverages such as tea, coffee and bottled water has changed. It was also assumed that there was no change in water source at individual workplaces.

REFERENCES Arbuckle, T.E., Hrudey, S.E., Krasner, S.W., Nuckols, J.R., Richardson, S.D., Singer, P., Mendola, P., Dodds, L., Weisel, C., Ashley, D.L., Froese, K.L., Pegram, R.A., Schultz, I.R., Reif, J., Bachand, A.M., Benoit, F.M., Lynberg, M., Poole, C. & Waller, K. 2002. Assessing exposure in epidemiologic studies to disinfection by-products in drinking water: report from an international workshop. Environmental Health Perspectives 110 (suppl. 1): 53–60.

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Bates, M.N., Rey, O.A., Biggs, M.L., Hopenhayn, C., Moore, L.E., Kalman, D., Steinmaus, C. & Smith, A.H. 2004. Case-control study of bladder cancer and exposure to arsenic in Argentina. American Journal of Epidemiology 159: 381–389. Canadian Ministry of National Health and Welfare 1981. Tapwater consumption in Canada. Document number 82-EHD-80. Public Affairs Directorate. Department of National Health and Welfare. Ottawa, Canada, 1981. Cantor, K.P., Hoover, R., Hartge, P., Mason, T.J. & Silverman, D.T. 1987. Bladder cancer, drinking water source, and tapwater consumption. A case-control study. Journal of the National Cancer Institute 79: 1269–1270. Colt, J.S., Baris, D., Clark, S.F., Ayotte, J.D., Ward, M., Nuckols, J.R., Cantor, K.P., Silverman, D.T. & Karagas, M. 2002. Sampling private wells at past homes to estimate arsenic exposure: a methodological study in New England. Journal of Exposure Analysis and Environmental Epidemiology 12: 329–334. Ershow, A.G. & Cantor, K.P. 1989. Total water and tapwater intake in the United States: population-based estimates of quantities and sources. National Cancer Institute Order #263-MD-810264, Bethesda, MD, 1989. Ershow, A.G., Brown, L.M. & Cantor, K.P. 1991. Intake of tapwater by pregnant and lactating women. American Journal of Public Health 81: 328–334. Foundation for Water Research. 1996. Tap water consumption in England and Wales: findings from the 1995 national survey. Report No. DWI0771, 1996. Focazio, M.J., Welch, A.H., Watkins, S.A., Helsel, D.R. & Horn, M.A. 2000. A retrospective analysis on the occurrence of arsenic in ground-water resources of the United States and limitations in drinking-water-supply characterizations: U.S. Geological Survay Water-Resources Investigations Report 99–4279, 21 pp. 2000. Hinkle, S.R. & Polette, D.J. 1999. Arsenic in ground water of the Willamette Basin, Oregon. Water Resources Investigations Report 98–4205: U.S. Geological Survay, Portland, OR, 28 pp. 1999. Karagas, M.R., Le, C.X., Morris, S., Blum, J., Lu, X., Spate, V., Carey, M., Stannard, V., Klaue, B. & Tosteson, T.D. 2001. Markers of low level arsenic exposure for evaluating human cancer risks in a US population. International Journal of Occupational Medicine and Environmental Health 14: 171–175. Levallois, P., Guévin, N., Gingras, S. Lévesque, B., Weber, J.-P. & Letarte, R. 1998. New patterns of drinkingwater consumption: results of a pilot study. The Science of the Total Environment 209: 233–241. Robertson, B., Forbes, A., Sinclair, M., Black, J., Veitch, M., Pilotto, L., Kirk, M. & Fairley, C.K. 2000. How well does a telephone questionnaire measure drinking water intake? Australian and New Zealand Journal of Public Health 24: 619–622. Roseberry, A.M. & Burmaster, D.E. 1992. Lognormal distribution for water intake by children and adults. Risk Analysis 12: 99–104. Ryan, P.B., Huet, N. & MacIntosh, D.L. 2000. Longitudinal investigation of exposure to arsenic, cadmium, and lead in drinking water. Environmental Health Perspectives 108: 731–735. Shimokura, H.H., Savitz, D.A. & Symanski, E. 1998. Assessment of water use for estimating exposure to tap water contaminants. Environmental Health Perspectives 106: 55–59. Steinmaus, C., Yaun, Y., Bates, M.N. & Smith, A.H. 2003. Case-control study of bladder cancer and drinking water in the Western United States. American Journal of Epidemiology 158: 1193–1201. Swan, S.H. & Waller, K. 1998. Disinfection by-products and adverse pregnancy outcomes: what is the agent and how should it be measured? Epidemiology 9: 479–481. USDA 1995. Food and nutrient intakes by individuals in the United States, 1 day, 1989–91. United States Department of Agriculture, Agricultural research Service, NFS Report No. 91–2, 1995. USEPA 1997. Exposure Factors Handbook. National centre for Environmental Assessment, Office of research and Development, Washington DC, USA. EPA\600\P-95\002Fabc, August 1997. USEPA 2000. Estimated per capita water ingestion in the United States. Office of Science and Technology, Office of water, Washington DC, USA. Williams, B.L., Florez, Y. & Pettygrove, S. 2001. Inter- and intra-ethnic variation in water intake, contact, and source estimates among Tucson residents: Implications for exposure analysis. Journal of Exposure Analysis and Environmental Epidemiology 11: 510–521.

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Arsenic bioaccumulation in a green algae and its subsequent recycling in soils of Bangladesh Imamul Huq, Afroza Bulbul & M.S. Choudhury Department of Soil, Water & Environment, University of Dhaka, Dhaka, Bangladesh

Shah Alam & Shigenao Kawai Department of Agro-Bioscience, Iwate University, Morioka, Japan

ABSTRACT: A green algae (Pithophora) collected from the botanical garden of Dhaka University were grown in vitro containing or not added arsenic (As). Arsenic application rates were 0.0, 0.05, 0.25, 0.5, and 1.0 mg/L. The algae were grown for 90 d, during which the algae received in increments the treatments of As at the specified concentration. After 90 d the algae were collected through filtration. The collected algae were dried, pulverized, and chemically analyzed for As. The treated algae were then added to a soil and were allowed to decompose over a period of 45 d. Soil samples were collected at intervals of 15, 30, and 45 d. The decomposition of algae to soil enrichment with As was studied. Presence of As in the algal growth medium substantially increased the cellular As content of algae and was contributory to soil enrichment by the element upon mineralization.

1

INTRODUCTION

Arsenic (As) in groundwater and its fate and transport in the environment have become matters of great concern in Bangladesh. Tube-well water extracted in Bangladesh from shallow aquifers is the primary source of drinking/cooking water for most of its population. Besides domestic use, huge quantities of water from shallow aquifer are also used for irrigation during the dry season. Since its detection in late 1993 in Bangladesh, much of the research works on As have focused on its presence in and exposure through drinking/cooking water (Yousuf et al. 2001). However, widespread use of groundwater for irrigation suggests that ingestion of irrigated crops could be another major exposure route for As. Another major concern is the phytotoxicity due to increased As in soil/water and its long-term impact on agricultural yield is another major concern (Ali et al. 2003). Limited studies have been conducted to assess the presence of As in the food chain (Meharg et al. 2001, Huq et al. 2001, Duxbury et al. 2002, Abedin et al. 2002). The country is heavily dependent on its agricultural sector for its gross domestic products (GDP). More than 80% of the population depends on agriculture for its livelihood. The agricultural sector employs about 90% of rural males as well as 80% of rural females of the country (BBS 1998). To become self-sufficient in food grain, Bangladesh government has encouraged production of High Yielding Variety (HYV) rice, which requires a large volume of irrigation water. The use of groundwater for irrigation has increased abruptly over the last couple of decades. About 86% of the total groundwater withdrawn is utilized in agricultural sector (BADC 2000). The use of As contaminated irrigation waters in Bangladesh may cause accumulation of As in rice and rice plants and this issue needs to be examined. Since rice is the staple food in this country, any adverse effects on nutrient content of rice due to arsenic contaminated irrigation water would enhance the malnutrition problem although As uptake and accumulation in rice plant from irrigation water may differ depending on cultivars used (Abedin et al. 2002). About 40% of total arable land of our country is now under irrigation facilities and more than 60% of this irrigation water come from groundwater which is extracted by deep 119

tube-well, shallow tube-well, and hand tube well (BBS 1998). Most groundwater used for irrigation in Bangladesh is contaminated with As (Khan et al. 1998). Twenty percent loss of crop (cereal) production due to high concentration of As (30 mg/kg) in plant body was reported by Davis & Coker (1979). Like other heavy metals, As is toxic to plants. In Bangladesh, Boro rice (dry season rice) is the major recipient of irrigation water. The rapid increase in Boro production is mainly attributed to the increase in the area under irrigation (and also use of high yielding varieties of rice). Currently, Boro rice accounts for about 37% of the total rice production in Bangladesh (BBS 1998). Besides Boro rice, wheat and a range of other crops and vegetables cultivated during dry season also need irrigation. Historically, the growth of algae in rice fields has been considered a natural fertilization process as decomposition of algae in rice fields adds nitrogen and other nutrients to the soil. The algae growing in the rice fields are supposed to take up among others, the As present in the water. Upon decomposition, the As thus accumulated will be mineralized and enrich the soil with the toxic element and it is possible that the subsequent crop takes it up. This is not a desirable situation. Keeping these views in mind, a laboratory experiment with green algae was set up with the objectives to study the bioaccumulation of As by algae from As contaminated water and to analyze the recycling of the bioaccumulated As.

2 2.1

MATERIALS AND METHODS Bioaccumulation of As

Seeds of green algae (Pithophora) of the Phytophthera family were collected from the Botanical garden of Dhaka University. A stock of the algae was produced by growing them in nutrient solution of the following composition: nitrogen (24 g/L N), potassium (2.21 g/L K), and phosphorus (8 g/L P) supplied from NH4Cl, KH2PO4, and CaHPO4, respectively (Smilde 1981). All salts used were of Analar grade. For the culture of algae in the laboratory, 5 treatments (0, 0.05, 0.25, 0.5, and 1.0 mg/L As in water) were chosen. The As was applied from sodium meta-arsenite. Each treatment was replicated 3 times. After each pot was filled with tap water and the nutrient solution, the As solutions of different concentrations were added to them. Then 20 ml of the green algae from the stock were transferred into each pot. The algae were cultured for 90 d receiving full sunlight during the day to get a substantial biomass. During this period, water containing As at the specified concentration were added. As such, the treatments received a total of 0, 2.5, 12.5, 25 and 50 mg As respectively for the five treatments over 90 d of growth. This was done to keep the water level in the pots at the same level as in the beginning of the experiment. After 90 d, the algae were harvested carefully from the pot by filtration and were allowed to dry in air. After drying the algae in air, these were oven-dried and dusted by a blender and the dust algae were latter used for mineralization process in soil. Results presented here are the averages of three individual replications. 2.2 Recycling of the elements in soil The soils used in the present study were collected from 0–150 mm depth. The soils were noncalcareous Grey Floodplain and represented the Dhamrai series, a Typic Haplaquept (mixed, nonacid, silty clay). The soil samples were collected from arable land and air dried in the laboratory. For hastening the drying it was exposed to sunlight. After drying in air, the larger aggregate were broken gently by crushing it in a wooden mortar and passed through a 2 mm sieve. Plastic pots of 250 mL sizes, washed with detergent and rinsed with distill water and dried in sunlight, were used for the incubation study. Each pot was filled with 100 g of dry soil and about 10 g harvested algae were added to each pot and the soil was properly submerged with sterilized water. The soils remained submerged with sterilized water for 45 d. All experiments were done in triplicates. Soils from each pot were collected at an interval of 15, 30, and 45 d. The soils were dried, crushed, and stored in a plastic bag for analysis. Results presented here are the averages of three individual replications. 120

Table 1. Some physical, physico-chemical, and chemical properties of soil used during the study.

2.3

Properties

Values

pH (soil:water ⫽ 1:2.5) Electrical conductivity (ECe) Particle size analysis Sand Silt Clay Texture Moisture content Organic carbon (OC) Organic matter (OM) Total nitrogen Total phosphorus Total potassium Total sulfur Available N Available P Available K Available S Arsenic (As)

6.28 0.07 ␮S 6.26% 48.68% 42.05% Silty clay 3.8% 0.8% 1.4% 0.12% 0.07% 0.32% 0.17% 58.24 mg/kg 3.024 mg/kg 0.2564 meq 42 mg/kg n.d.

Analysis of algal samples

The algal samples were analyzed at the time of collection and after harvest for As (Jackson 1973). Arsenic was analyzed by hydride generation atomic absorption spectrophotometric (HG-AAS) technique. Arsenic in algae and in soil was extracted following the method of Portman & Riley (1964). Arsenic in algae at the time of collection was not in the detectable range. 2.4

Analysis of soil sample

Soils were also analyzed before the set up of the experiments to see the nutrients status at the time of collection. The results are presented in Table 1. Standard methods were followed for the routine analyses of soil. Arsenic in soil was analyzed as described for algae.

3 3.1

RESULTS AND DISCUSSION Bioaccumulation of As in algae

The bioaccumulation of As increased gradually with increasing As concentration in water (Fig. 1). When As was absent in the solution the algae did not show detectable quantity of As. However, even at 0.05 mg/L As in water application rate (2.5 mg As) corresponding As in algae was 11.14 mg/kg and at 1.0 mg/L As treated water (50.0 mg As) the algae had accumulated 53.19 mg/kg As. Algae possess the ability to take up toxic heavy metals from the environment, resulting in higher concentration than that in the surrounding water (Megharaja et al. 2003). Bioaccumulation studies reveal the accumulation of contaminant in the organism via food or water containing the contaminant (Megharaja et al. 2003). Plants take up As3⫹ passively with water, but As5⫹ is actively taken up by some algal species (Andreae et al. 1983). Similar high accumulation of As by a fern has been reported by Ma et al. (2001). The regression analysis between applied As and As in algae showed that the relationship was positive (Y ⫽ 6.251 ⫹ 49.287x) and highly significant (r ⫽ 0.9578**). 121

As content in algae, mg kg-1

70 60 50 40 30 20 10 0 0

Figure 1.

2.5 12.5 25.0 Arsenic applied in water (mg)

50.0

Content of As in algae as affected by applied As in water. 35 As in soil after 15 d

30

As in soil after 30 d As in soil after 45 d

As in soil, mg/kg

25 20 15 10 5 0

To

T1

T2

T3

T4

To

T1

T2

T3

T4

To

T1

T2

T3

T4

Figure 2. Mineralized As in soil from algae treated with different concentration of As (T0, T1, T2, T3 & T4 are 0, 0.05, 0.25, 0.5 & 1.0 mg As/L).

3.2

Recycling of As in soil

The release of As through mineralization of algae was studied at different phases of incubation (Fig. 2). It is apparent that the algae growing on As-contaminated water and bioaccumulating it subsequently released the element in soil during the process of mineralization. The release pattern however, differed with days of incubation. The release of As was maximum after 30 d of incubation followed by a decrease after 45 d of incubation. The release of As at 15 d was the maximum in soils receiving algae treated at the rate of 1.0 mg As/L. After 30 d of incubation the picture was little bit different. The maximum As release was in the soil receiving algae treated with 0.5 mg/L As. It needs to be mentioned here that the background level of As in soil was less than the detectable limit of 2 ␮g/L. Hence, it is clear that algae growing in environments of As contamination accumulate it and subsequently release it to the soil. The analysis of variance (ANOVA) indicated that both algal As content as well as days of incubation had significant positive contribution to As release in soil (F for As in algae 13.054** and for days of incubation 10.569**, LSD(0.95) being 5.96). A large fraction of this recycled As is not likely to be washed out by flood or rainwater in oxidized condition due to its high affinity to iron (Fe), manganese (Mn), aluminum (Al), and other elements in soils (Kabata-Pendias & Pendias 2001). As a result, a cumulative accumulation of As in surface soils is expected and some recent data show As contents as high as 83 mg/kg in the top soil of Bangladesh (Huq et al. 2003, Alam & Sattar 2000, Ullah 1998). In general, higher As contents have been reported in the top layer of soils (Huq et al. 2001). The increased As content after 30 d 122

of incubation could be due to the maximum mineralization of the added algae. The reduction in As release after 45 d could be a phenomenon of As fixation or conversion through biomethylation. This result also indicates that rice soils irrigated with As contaminated water, where algal growth is a common occurrence, will accumulate As and will be taken up by the crop grown on such soils. Arsenic concentration in algae indicated positive correlation with the As content in the soil for all the three different phases. Correlation coefficient was high (r ⫽ 0.91) for the 1st phase, whereas the values of r for 2nd and 3rd phases were 0.41 and 0.60, respectively. It may be concluded from the present investigation that rice fields with algal growth as a natural phenomenon and receiving irrigation from As contaminated groundwater, are likely to have a build up of As through nutrient recycling process. This is obviously a major concern vis-a-vis soil pollution and contamination of the food chain through water-soil-crop transfer of contaminants. This demands special attention so far the irrigated rice culture is concerned.

REFERENCES Abedin, M.J., Cresser, M.S., Meharg, A.A., Feldmann, J. & Cotter-Howells, J. 2002. Arsenic accumulation and metabolism in rice. Env. Sci. Technol. 36: 962–968. Alam, M.B. & Sattar, M.A. 2000. Assessment of arsenic contamination in soils and waters in some areas of Bangladesh. Water Sci. Technol. 42: 185–193. Ali, M., Ashraf, A.B.M.B., Jalil, M.A., Delwar Hossain, M., Ahmed, M.F., Al Masud, A., Kamruzzaman, M. & Rahman, M.A. 2003. Arsenic in plant soil Environment in Bangladesh. In M.F. Ahmed, A.M. Ashraf, & A. Zafar (eds.): Fate of Arsenic in the Environment, p. 85–112, BUET-UNU, Dhaka. Andreae, M.O., Byod, J.J. & Forelicj, P.N. 1983. Arsenic, Antimony, Germanium and Tin in the Jyo estuary, Portugal-modeling a polluted estuary. Environ. Sci. Tech. 17: 731–737. BADC 2002. Survey Report on Irrigation Equipment and Irrigated area in Boro/2001 season, Bangladesh Agricultural Development Corporation, March 2002. BBS 1996. Statistical Yearbook of Bangladesh, Bangladesh Bureau of Statistics, Statistics Division, Ministry of Planning, People’s Republic of Bangladesh. p. 10. Davis, R.D. & Coker, E.G. 1979. Proc. Int. Conf. Management and Control of Heavy Metals in the Environment. 553 CEP Consultants Ltd. Edinburg, UK. Duxbury, J.M., Mayer, A.B., Lauren, J.G. & Hassan, N. 2002. Arsenic content of rice in Bangladesh and Impacts on rice productivity. 4th Annual Conference on Arsenic contamination in groundwater in Bangladesh: Cause, Effect and Remedy, Dhaka, 12–13 January, 2002. Huq, S.M.I., Ahmed, K.M., Sultana, N. & Naidu, R. 2001. Extensive arsenic contamination of groundwater and soils of Bangladesh. In Arsenic in the Asia-Pacific Region Abstracts, p. 94–96, CSIRO-ACIAR, Australia. Huq, S.M.I. Rahman, A., Sultana, N. & Naidu, R. 2003. Extent and severity of arsenic contamination in soils of Bangladesh. In M.F. Ahmed, A.M. Ashraf, & A. Zafar (eds.): Fate of Arsenic in the Environment, p. 69–84, BUET-UNU, Dhaka. Jackson, M.L. 1973. Soil Chemical Analysis. Prentice Hall of India Pvt. Ltd., New Delhi. Kabata-Pendias, A. & Pendias, H. 2001. Trace Elements in Soils and Plants. 3rd Edition. CRC Press, Washington DC, p. 331. Khan, A.W., Ahmad, S.A., Sayed, M.H.S.U., Hadi, S.A., Khan, M.H., Jalil, M.A., Ahmed, R. & Faruque, M.H. 1998. Arsenic contamination in ground water and its effect on human health with particular reference to Bangladesh. Abstract Volume, International Conference on Arsenic Pollution of Groundwater in Bangladesh. 8–12 February 1998, p. 109. Ma, L.Q., Komar, K.M., Tu, C., Zhang, W., Cai, Y. & Kennelley, E.D. 2001. A fern that hyper accumulates arsenic. Nature, 409: 579. Megharaj, M., Raagusa, S.R. & Naidu, R. 2003. Metal-algae interactions: Implications of Bioavailability. In R. Naidu, V.V.S.R. Gupta, S. Rogers, R.S. Kukana, N.S. Bolan & D.C.Adriano (eds.), Bioavailability, Toxicity and Risk Relationships in Ecosystems, 109-144. Science Publishers, Inc. Enfield (NH), USA. Meharg, A.A., Abedin, M.J., Rahman, M.M., Feldmann, J., Cotter Howells, J. & Cresser, M.S. 2001. Arsenic uptake and metabolism in Bangladesh rice varieties. Arsenic in the Asia-Pacific Region Abstracts, p. 45–46, CSIRO-ACIAR, Australia. Portman, J.E. & Riley, J.P. 1964. Determination of arsenic in seawater, marine plants and silicate and carbonate sediments. Anal. Chim. Acta. 31: 509–519.

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Smilde, K.W. 1981. Heavy metal accumulation in crops grown on sewage sludge amended with metal salts. Plant Soil, 62: 3–14. Ullah, S.M. 1998. Arsenic contamination of groundwater and irrigated soils in Bangladesh, International Conference on Arsenic Pollution of Groundwater in Bangladesh: Causes, Effects and Remedies, 3–5 January, Dhaka, p. 133. Yousuf, J., Zaman, Q.Q., Das, R., Islam, S., Salim, S., Morshed, M. & Roy. S. 2001. Community management of arsenic patients in Bangladesh. In Arsenic in the Asia-Pacific Region Abstracts, p. 65, CSIRO-ACIAR, Australia.

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Environmental behavior of arsenic in a mining zone: Zimapán, Mexico M.A. Armienta, R. Rodríguez, O. Cruz, A. Aguayo, N. Ceniceros Instituto de Geofísica, UNAM, Circuito exterior C.U., México D.F., Mexico

G. Villaseñor Instituto de Geología, UNAM, Circuito exterior C.U., México D.F.

L.K. Ongley Oak Hill High School, Sabatus, USA

H. Mango Dept. of Natural Sciences, Castleton State College, Castleton, USA

ABSTRACT: The distribution and fate of arsenic in various environmental reservoirs have been studied in Zimapán, México. Natural arsenic in groundwater is released by oxidation and dissolution of arsenic-bearing minerals. Mining wastes (tailings and smelter particulates) pollute some of the shallow wells. Tailings contain high concentration of As. However, As mobility is limited by oxidation-reduction, sorption, and formation of secondary minerals. Sequential extraction showed that As is mainly retained in the residual and Fe/Al oxyhydroxides fractions. Arsenic occurs also in beudantite and K-jarosite. High arsenic concentrations are found in soils and river sediments near tailings, slag piles and mineralized areas. Due to the semi-arid climate, few crops are grown in this area. Nevertheless, wild flora has grown on tailing and slag piles, and As-rich soils. Arsenic is absorbed from these As-rich substrates and translocated to the stem, leaves, flowers and fruits. The concentrations and fate of arsenic varies depending on the arsenic source.

1

INTRODUCTION

Arsenic concentrations above the drinking water standard have been found in the groundwater of various parts of México (Armienta 2003). Some of this pollution has occurred in mineralized areas. Mining has been an important economic activity in México since the XVI Century. Many towns were founded and developed around mines. As a result of the ore extraction and processing, wastes were produced and accumulated in these towns. The accumulated residues have the potential to release metals and metalloids that pollute the environment. Furthermore, high concentration levels of toxic metals may be present naturally in mining zones (Runnels et al. 1992). Arsenic minerals are widespread in México which is the fifth largest As producer in the world. Although health effects of arsenic-polluted groundwater do not represent a major nation-wide problem, in locations like Comarca Lagunera it became an endemic problem (Cebrián et al. 1983). Besides, about 10% of milk samples collected from dairy farms at Comarca Lagunera were reported to contain arsenic over the permitted level (Rosas et al. 1999). More than 60% of the drinking water in México comes from groundwater. It is thus important to be able to identify the sources and understand the fate of arsenic in all the exploited aquifers in México. Zimapán is a low-income community with nearly 15,000 inhabitants. Groundwater is the only drinking water source for residents of this region. This mining area has been active for more than 400 years. The population living in the urban part of Zimapán has consumed water containing As 125

160

R2 = 0.444

140 SO4 (mg/L)

120 100 80 60 40 20 0 0

Figure 1.

0.2

0.4

0.6 As (mg/L)

0.8

1

1.2

Arsenic vs sulfate concentrations in deep wells in limestone aquifer.

varying from 0.19 mg/L to 0.65 mg/L (average 0.38 mg/L) for more than twelve years. In addition, some residents in the outskirts of the urban area were supplied with water of higher As content for these years. Health effects related to As intake have been detected in Zimapán inhabitants (Armienta et al. 1997a).

2

ARSENIC IN GROUNDWATER

Zimapán, located in a semi-arid zone, in the central part of México, is one of many mining districts in the country. Silver, Pb and Zn are currently exploited in Zimapán. The ore occurs as skarn and chimney/manto mineralization. The main minerals occur as massive sulfide ores: pyrite, pyrrhotite, sphalerite, galena, chalcopyrite, arsenopyrite, tetrahedrite and lead sulfosalts (Villaseñor et al. 1987). In addition to arsenopyrite, several other arsenic minerals also occur: scorodite (FeAsO4⭈2H2O), lolingite (FeAs2), tennantite ((Cu,Fe)12As4S13), adamite (Zn2(AsO4)(OH), mimetite (PbS(AsO4)3Cl) and hidalgoite (PbAl3(AsO4)(SO4)(OH)6). A complex aquifer system underlies the Zimapán basin. A deep fractured aquifer is developed in Cretaceous limestones (Soyatal and Tamaulipas formations). A granular shallow aquifer exists within Quaternary alluvium and Tertiary volcanic rocks in the center of the basin. A volcanic aquifer is developed in the eastern part of the valley (Armienta et al. 1997b). The more productive wells (more than 30 L/s) drilled in the fractured limestone aquifer also contain the highest As concentrations. Those wells are located in the proximity of intrusive bodies and dikes where As minerals are found (Armienta et al. 2001). Groundwater with low As concentration (from non-detectable up to 0.05 mg/L) is found in the volcanic aquifer, which has low productivity (less than 10 L/s). Arsenopyrite oxidation and scorodite dissolution have produced arsenic contamination in some of the deep wells (up to 180 m depth) drilled in limestones. A correlation may be observed between As and sulfate in those wells (Fig. 1). Other processes (mainly FeS dissolution and oxidation of As-rich pyrite) may release As and sulfate to those wells. FeAsS oxidation is reflected also in the relatively lower Eh, higher temperature (Armienta et al. 2001) and lower HCO3/SO42⫺ ratios recorded in polluted wells. Tailing piles oxidation has polluted nearby shallow wells. These wells had higher sulfate contents than deep contaminated wells. A mixing line among As-free water within the shallow aquifer and tailings leachate reflects the interaction between tailings and shallow groundwater (Fig. 2). Sulfate ions result from oxidation of various sulfide minerals within the tailings, mainly pyrite, pyrrothite, arsenopyrite, chalcopyrite, and sphalerite (Romero et al. 2004a). The correlation between arsenic and sulfate shows the influence of the oxidation of the tailings. 126

2000 1800 1600 SO4 (mg/L)

1400

R2 = 0.9026

1200 1000 800 600 400 200 0 0

0.2

0.4 0.6 As (mg/L)

0.8

1

Figure 2. Arsenic vs sulfate in shallow wells. Polluted wells are located next to tailing piles. Open circle corresponds to a tailings’ leachate composition. High sulfate concentrations reflect tailings’ oxidation. 200 180 160

R2 = 0.7642

SO4 (mg/L)

140 120 100 80 60 40 20 0 0

0.01

0.02

0.03

0.04 0.05 As (mg/L)

0.06

0.07

0.08

Figure 3. Arsenic vs sulfate concentration in shallow wells away from tailings. Arsenic contamination results from natural mineralization.

Lower As concentrations were measured in shallow wells within the Soyatal limestone, in a small zone near mineralized dikes away from the influence of the tailings (Fig. 3). A good correlation was observed between As and SO42⫺ in these wells further indicating that As results from oxidation of arsenic-bearing sulfides. Arsenic is mobilized by different mechanisms in the shallow granular aquifer and in the deep fractured limestone aquifer. Arsenic in the deep aquifer is transported through fractures, with a higher water-rock ratio. Arsenic in shallow wells flows as a function of sediment permeability in the granular aquifer, resulting in a lower water-rock ratio. Released As may be retained on calcite, iron oxyhydroxides and clays in the aquifer matrix (Romero et al. 2004b).

3

ARSENIC IN TAILINGS

Old and some recent tailing piles are found along the Zimapán town margins. The main mineralogy includes quartz, calcite, gypsum, arsenopyrite, and jarosite. Sulfates in water leachates ranged 127

from 512 mg/kg in unoxidized tailings to 2750 mg/kg in oxidized tailings. The influence of calcite on the physico-chemical characteristics of these wastes is reflected in the near neutral pH values of most samples (from 6.4 to 7.6). A low pH (3.3) was measured in one oxidized pile whose mineralogy indicates calcite consumption and formation of secondary minerals such as gypsum and jarosite. High As contents have been measured in all tailing piles (up to 22,000 mg/kg) in spite of their different oxidation degress (Armienta & Rodríguez 1996). Nevertheless, As concentration in shallow wells next to tailings reached only 0.44 mg/L. These relatively low As concentrations in groundwater may be explained by As retention processes occurring within the tailings. Mendez & Armienta (2003) determined, through sequential extraction, that arsenic is mostly associated with the low mobility Fe and Al oxyhydroxides, and other residual fractions in four tailing piles in Zimapán. Arsenic associated with the most mobile soluble and exchangeable fractions was low in most tailings (from 3 to 9%). However, this percentage reached 34% in one of the oxidized piles that at the same time had a lower proportion in the Fe and Al fraction . This indicates that only part of the arsenic may be available to be mobilized. The formation of arsenic-bearing secondary minerals, mainly beudantite and jarosite, and As sorption onto iron oxyhydroxides are also controlling As dispersion (Romero et al. 2004a). These interactions have limited arsenic dispersion preventing higher concentrations than those observed in shallow wells next to tailings. Tailing piles occur on top of alluvial material of high permeability; this has allowed leachate percolation. Differences between tailings leachate concentration measured by Ongley et al. (2001) and groundwater As in nearby wells indicates that water-rock interactions are playing an important role in As retention.

4

ARSENIC IN SOILS AND SEDIMENTS

Arsenic enrichment in soils next to tailing piles, former smelters and mineralized zones has occurred at Zimapán. Superficial soils near tailings contain up to 2580 mg/kg of total As and 7.3 mg/kg of soluble As. Total and soluble As kept high values in 30 cm depth samples (up to 1070 mg/kg of total As, and 8.4 mg/kg of soluble As) (Armienta & Rodríguez 1995). Arsenic concentrations up to 4200 mg/kg were measured in soils near former smelters or smelter wastes. Soluble As increases with depth in some sites from 5 mg/kg on the surface to 14.3 mg/kg at 50 cm depth. Settling of As-rich smelter particulates, wind dispersion of solid tailings, and leachate runoff from tailings has contaminated soils around smelters and tailings. In addition, natural subsurface As mineralization has resulted on As concentration as high as 768 mg/kg in mineralized zones. Soils more than 4000 m from tailings and slags generally had less than 40 mg/kg As (Ongley et al. 2003). Percolation of water through polluted soils has also increased As water content in shallow wells near former smelters. Sediments of the Tolimán river are also contaminated with arsenic. This non-perennial river is the only surficial water body in the Zimapán basin. It flows next to tailings and reaches the mine zone. The highest concentrations (up to 6575 mg/kg) were measured in front of the tailings. Arsenic content also increases in the mine area, reaching 5148 mg/kg (García et al. 2001). Arsenic is mainly associated with Fe and Mn oxyhydroxides and residual fractions in sediments. Arsenopyrite was identified in As-rich sediments influenced by tailings and in the mineralized zone (García 1997). Higher As contents corresponded to lower pH values. Oxidation of arsenopyrite and other sulfide minerals may decrease the pH and increase the arsenic concentration. Released As may then be partly retained onto Fe and Mn oxyhydroxides.

5

ARSENIC IN PLANTS

Arsenic was analyzed in samples from mesquite (Prosopis laevigata), huizache (Acacia farnesiana), and pepper tree (Schinus molle), all of which are naturally growing in mining wastes (Fig. 4). The highest concentration was measured in a mesquite root (1400 mg/kg) located 128

Figure 4.

A mesquite tree growing on a tailings pile.

on a non-oxidized tailing pile. Much lower concentration was found in leaves of the same tree (66 mg/kg). These concentrations are much higher than those found in mesquite growing on agar solution with 5 ppm As2O3 (Aldrich et al. 2002). A similar As content was measured in the root of the fern Pteris vittata. This fern has been found to be extremely efficient in extracting arsenic from soils. Fern specimens growing on a contaminated soil with chromate copper arsenate contained 1442–7526 ppm, and much higher concentrations (up to 21290 ppm) were accumulated by this fern in plants when grown over artificially contaminated soils containing 500 ppm As (Ma et al. 2001). A huizache tree growing on a slag pile also had high As content (102 mg/kg in leaves, and 119 in stems). Arsenic concentration in a pepper tree on a slag pile was 99 mg/kg in the stem, 115 mg/kg in the leaves and 62 mg/kg in the flowers. Lower As concentrations were measured in pods (up to 30 mg/kg). These plants are not edible; however, mesquite and huizache leaflets and pods are consumed by goats, which are the most important livestock in the area. In addition, cultivated areas cover a small proportion of the total Zimapán region (about 3%). Plants may therefore indirectly be another source of As for Zimapán inhabitants. 6

CONCLUSIONS

Arsenic is widely distributed in Zimapán. Mining wastes and natural arsenic-bearing mineralization release As to water, soils and plants. Geochemical processes create differences among these sources. The highest As concentrations in groundwater are produced by natural processes, primarily the oxidation and dissolution of As-bearing minerals. This process takes place mainly near mineralized zones. Arsenic from this source may be transported long distances in the deep fractured aquifer. Mining wastes have polluted shallow wells near the residues. Once in the shallow aquifer, As transport may be retarded by water-rock interactions. Complex processes occur within tailings that release and retain arsenic. Although total As concentrations up to 22,000 mg/kg have been measured in tailings, only a low proportion (less than 10% for most tailing piles) is easily available to the environment. Soils and sediments are also enriched in As due to the presence of tailings, smelter particulates and general arsenic mineralization. Water infiltration through these substances pollute some shallow wells. Wild plants growing on mining residues absorb arsenic which is then mainly concentrated in roots. However, As is translocated to stems, leaves, flowers and pods and may enter the food chain by being consumed by goats. 129

Understanding of As behavior in mining areas requires comprehensive studies to identify the sources of As and their contribution to As contamination. This may allow for the development of remediation alternatives that at the same time allow economic and social development of communities relying on mining, such as Zimapán. ACKNOWLEDGEMENTS The authors thank the financial support given by grant 017 of the National Council of Science and Technology, CONACyT, -SEMARNAT Mexico, by UNAM, and by the National Science Foundation (USA). REFERENCES Aldrich, M.V., Parsons J.G. & Gardea-Torresdey, J.L. 2002. Arsenic(V) and (III) uptake by the desert plant specites mesquite (Prosopis spp.), Application of waste remediation technologies to agricultural contamination of water resources, July 30–August 1, 2002, Kansas City, Missouri, USA. Armienta, M.A. 2003. Arsenic Groundwater Pollution in Mexico. Medical Geology Newsletter 6: 4–6. Armienta, M.A. & Rodríguez, R. 1995. Evaluación del riesgo ambiental debido a la presencia de arsénico en Zimapán, Hidalgo. Memoria Final, Fundación MAPFRE, IGF, UNAM, 42 pp. Armienta, M.A. & Rodríguez, R. 1996. Arsénico en el Valle de Zimapán, México: Problemática Ambiental, Revista MAPFRE Seguridad, Madrid, España. 63: 33–43. Armienta, M.A., Rodríguez, R. & Cruz, O. 1997a. Arsenic content in hair of people exposed to natural arsenic polluted groundwater at Zimapán, México. Bulletin of Environmental Contamination and Toxicology 59: 583–589. Armienta, M.A, Rodríguez, R., Aguayo, A., Ceniceros, N., Villaseñor, G. & Cruz, O. 1997b. Arsenic contamination of groundwater at Zimapán, Mexico. Hydrogeology Journal 5: 39–46. Armienta, M.A., Villaseñor G., Rodríguez R., Ongley L.K. & Mango, H. 2001. The role of arsenic-bearing rocks in groundwater pollution at Zimapán Valley, México. Environmental Geology 40: 571–581. Cebrián M.E., Albores A., Aguilar M. & Blakely, E. 1983. Chronic arsenic poisoning in the north of Mexico. Human Toxicology 2: 121–133. García E.A. 1997. Distribución y especiación de arsénico en sedimentos fluviales del río Tolimán en Zimapán, Hgo. M. Sc. Thesis, UNAM, México D.F., 81 pp. García A., Armienta M.A. & Cruz O. 2001. Sources, distribution and fate of arsenic along the Tolimán river, Zimapán, Mexico. Red Book IAHS Publ. No. 266, UK, pp 57–64. Ma l.Q., Komar K.M., Tu C., Zhang W., Cai Y. & Kenneley, E.D. 2001. A fern that hyperaccumulates arsenic. Nature 409: 579. Méndez M. & Armienta, M.A. 2003. Arsenic Phase Distribution in Zimapán Mine Tailings, Mexico. Geofísica Internacional 42: 131–140. Ongley L.K., Armienta, M.A., Heggeman, K., Lathrop, A.S., Mango, H., Miller, W. & Pickelner, S. 2001. Arsenic removal from contaminated water by the Soyatal Formation, Zimapán District, Mexico – a potential low-cost low-tech remediation system. Geochemistry; Exploration, Environment, Analysis 1: 23–31. Ongley, L.K., Armienta, M.A. & Mango, H. 2003. Concentrations of heavy metals in soil, Zimapan, México. Journal de Physique IV 107: 983–986. Romero F.M., Armienta M.A., Villaseñor, G. & González, J.L. 2004. Mineralogical constraints on the mobility of arsenic in tailings from Zimapán, Hidalgo, Mexico. International Journal of Environment and Pollution (accepted). Romero F.M., Armienta M.A. & Carrillo-Chávez, A. 2004. Arsenic Sorption by carbonate-rich aquifer material, a control on arsenic mobility at Zimapán, México. Archives of Environmental Contamination and Toxicology (In Press). Rosas I., Belmont R., Armienta M.A. & Baez, A. 1999. Distribution of Arsenic Levels in Dairy Farms at Comarca Lagunera, Mexico. Water, Air & Soil Pollution 112: 133–149. Runnels D.D., Shepherd T.A. & Angino, E.E. 1992. Metals in water. Determining natural background concentrations in mineralized areas. Environmental Science and Technology 26: 2316–2323 Villaseñor C.M.G., Gomez-Caballero, J.A., Medina de la Paz, J.L. & Lozano, R. 1987. Boulangerita de chimenea Las Animas, Zimapán, Hidalgo: mineralogía y metalogenia. Boletín de Mineralogía 3: 1–30.

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Section 3: Arsenic biogeochemistry in groundwater

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Natural enrichment of arsenic in groundwaters of Brahmanbaria district, Bangladesh: geochemistry, speciation modeling and multivariate statistics Ondra Sracek Institute of Geological Sciences, Faculty of Science, Masaryk University, Brno, Czech Republic

Prosun Bhattacharya, Mattias von Brömssen, Gunnar Jacks Groundwater Arsenic Research Group, Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden

Kazi Matin Ahmed Department of Geology, University of Dhaka, Dhaka, Bangladesh

ABSTRACT: Groundwater with geogenic arsenic enrichment is commonly encountered in the Holocene sedimentary aquifers of the Bengal Delta Plain (BDP). The present study was carried out in Brahmanbaria district, covering an area of 18 km2 in northeastern Bangladesh. The Chandina Formation is the main hydrostratigraphic unit of the area, which comprises silt and clay with high content of organic matter. Dissolved arsenic concentrations in groundwater are high, reaching ⬎400 ␮g/L in some wells. Groundwater is reducing with general lack of detectable dissolved oxygen (DO) and contains low concentrations of nitrate and sulfate. Concentrations of dissolved Fe are high, which is in general in agreement with the reductive dissolution of ferric oxide and hydroxide hypothesis. Results of speciation modeling indicated the possibility of precipitation of siderite, and to less extent, vivianite for many samples. The log PCO2 values were extremely high (⬎⫺1.0 atm), suggesting production of CO2 in redox reactions involving the organic matter in the sediments. Redox potential values calculated on the basis of different redox couples and field Eh measurement indicated redox disequilibrium. Hierarchical cluster analysis (HCA) performed in paired groups mode using the program PAST indicated highest degree of similarity among redoxsensitive elements NO3, Mn, Fe, PO4, SO4, As, and pH. Na and Cl form a distinct group, which indicate the influence of sea water. Bicarbonate generated in several redox reactions and carbonate dissolution was linked to almost all parameters and this holds even more for the electrical conductivity (EC). Principal components analysis (PCA) yielded Principal Component 1 (PC1) corresponding to sea water, and Principal Component 2 (PC2) corresponding to redox reactions with generally high arsenic concentrations. In summary, combination of speciation modeling and multivariate statistics proved to be useful in testing of conceptual model of geochemical evolution of arsenicrich groundwater.

1

INTRODUCTION

Natural arsenic in concentrations above the safe drinking water limits of World Health Organization (10 ␮g/L; WHO 2001), and above the national drinking water standard (50 ␮g/L) is present in groundwater of the Bengal Delta Plain (BDP) in many districts of Bangladesh (Mukherjee & Bhattacharya 2001, Smedley & Kinniburgh 2002, Ahmed et al. 2004). The source of arsenic is geogenic and is related to the sediments deposited by the rivers Ganges (Padma), Brahmaputra 133

(referred as Jamuna in Bangladesh) and Meghna (Nickson et al. 2000, Bhattacharya et al. 2002a,b). Arsenic contaminated groundwater is common in the aquifers of alluvial lowlands, comprising the floodplains of Padma and Brahmaputra (Jamuna) rivers, and also the Ganges Delta. In this paper, we present applications of geochemical modeling and multivariate statistics in development and support of conceptual model of arsenic behavior.

2

GEOLOGICAL SETTING

The BDP is a large sedimentary basin drained by the Ganges, Brahmaputra (Jamuna) and Meghna (GBM) rivers (Fig. 1). Huge amounts of sediments have been transported and converged at the lower reaches, forming the pro-grading delta at the head of the Bay of Bengal. In general, two broad physiographic units characterize the BDP – elevated Pleistocene Terraces such as the Barind and Madhupur tracts, floored with thick surficial oxidized clay and silty clay deposits, and the Holocene lowlands. The Holocene lowlands include piedmont plains, flood plains, delta plains and coastal plains (Umitsu 1987 and 1993, Brammer 1996, Ahmed et al. 2004). The area of present investigation is located in the Meghna Deltaic Plain comprising coarse-grained channel-fill deposits and fine grained overbank deposits. During the late Holocene period, in several parts of the BDP, sediments were deposited in marshy environments, as evidenced by occurrence of continuous layers of peat (Umitsu 1993, Ravenscroft et al. 2001, Ahmed et al. 2004). The present study was carried out in an area of 18 km2 covering the Ashuganj and Brahmanbaria Sadar Upazilas (sub-districts) in Brahmanbaria district in eastern Bangladesh (Fig. 2). The Chandina Deltaic Plain (CDP), the major physiographic units covering the study area (Bakr 1977), is generally flat and occurring at relatively higher levels than the surrounding floodplains. The

Figure 1. Map of Bangladesh showing the network of the rivers Ganges (Padma), Jamuna (Brahmaputra) and Meghna rivers and the location of Brahmanbaria area. The major geomorphic domains, Barind and Madhupur tracts of Pleistocene age (lighter tone) are seen distinctly within the vast tract of Holocene alluvium. (Resolution: 625 meters; MODIS Data Type: MODIS-PFM; MODIS Band Combination: 1, 4, 3) (Source map: http://modis.gsfc.nasa.gov/MODIS/IMAGE_GALLERY/MODIS1000027_md.jpg).

134

sediments of the CDP are composed of silt, silty loam, silty clay belonging to the Chandina Formation. The Chandina Formation is overlain by the Meghna alluvium and underlain by the Pleistocene Madhupur Clay and Pliocene Dupi Tila Formation. Neotectonic uplifts and course shifting of the Meghna and Old Brahmaputra rivers have influenced sedimentation in this area, particularly during the late Holocene time. The source terrains of most of these sediments were located in the areas around the Shillong Plateau in the north and Tripura Hills on the east. A thick sequence of fine to very fine Holocene sediments overlies the Pleistocene Madhupur Clay and Pliocene Dupi Tila sediments. The Holocene sediments are generally gray in color and contain high amount of organic matter while underlying older sediments are characterized by reddish-brown, light brown and yellowish brown color and contain only low amounts of organic matter (Table 1). Groundwater from the Holocene sandy sedimentary aquifer is extracted by shallow hand tube wells. Water levels in the Holocene aquifers fluctuate with annual recharge/discharge conditions, with a maximum depth of 5–7 m bgs in pre-monsoon months. During the monsoon most of the area if flooded and the groundwater level reaches the ground surface. The multiple aquifer system

Figure 2. Geological map of a part of the Chandina Delta Plain (CDP) in the Brahmanbaria district, Bangladesh showing the location of the groundwater sampling. Table 1.

Hydrostratigraphy of the study area.

Age

Unit

Predominant Lithology

Recent/Holocene

Meghna Alluvium

Grey clay, silt and fine sand

Late Pleistocene Early Pleistocene Pliocene

Chandina Formation Madhupur Clay Dupi Tila Sands

Hydrogeological characteristics

Upper unconfined aquifer, arsenic rich Grey silt, silty loam, silty clay Aquitard Reddish brown clay Aquitard Yellowish brown medium to Lower aquifer, low in arsenic fine sand

135

in this area is characterized by variable hydraulic conductivity and water quality. Water quality is often good except for occurrences of pockets of brackish water, remnants of paleo-seawater. Occurrences of biogenic methane gas have also been reported from a number of places with the BDP and particularly in the vicinity of the study area (Ahmed et al. 1998, Ravenscroft et al. 2001). Arsenic and iron concentrations are frequently high in the Holocene aquifers. However, their concentrations are significantly lower in the underlying Dupi Tila aquifers (BGS & DPHE 2001, van Geen et al. 2003). This deeper aquifer probably receives recharge at their outcrops in the Tippera Hills region, outside the political boundary of Bangladesh. 3

MATERIALS AND METHODS

Samples of groundwater were collected during late November, 2000 from 30 domestic and governmental tube wells placed at varying depths of 18–150 m (Fig. 2). Parameters like pH, redox potential (Eh), temperature, and electrical conductivity (EC) were taken in the field. The pH was measured using a Radiometer Copenhagen PHM 80 instrument using a combination electrode (pH C2401-7). The Eh was measured in a flow-through cell using a combined platinum electrode (MC408Pt) equipped with a calomel reference cell. Samples collected for analyses included: (a) filtered (using Sartorius 0.45 ␮m online filters) for major anions; (b) filtered and acidified with suprapure HNO3 (14 M) for the cations and other trace elements including arsenic (Bhattacharya et al. 2002b). Arsenic speciation was performed with Disposable Cartridges(r) (MetalSoft Center, PA) in the field, Meng et al. (2001). The cartridges adsorb As(V), but allows As(III) to pass through. Sulfide was precipitated in the field by addition of Zn acetate. Major anions, Cl⫺, and SO 42⫺ were analyzed in filtered water samples, with a Dionex DX-120 ion chromatograph with an 3⫺ IonPac As14 column. NO ⫺ 3 -N and PO 4 -P was analyzed spectrophotometrically with a Tecator ) was analyzed spectrophotometrically with a Tecator Aquatec Aquatec 5400. Ammonium (NH⫹ 4 5400 at 540 nm wavelength. The major cations (Ca, Mg, Na and K) and minor and trace elements (Fe, Mn, As) were analyzed by inductively coupled plasma (ICP) emission spectrometry (Varian Vista-PRO Simultaneous ICP-OES) at Stockholm University. Dissolved organic carbon (DOC) in the water samples were determined on a Shimadzu 5000 TOC analyzer (0.5 mg/L detection limit with a precision of ⫾10% at the detection limit. Speciation modeling was performed by program PHREEQC (Parkhurst 1995). Thermodynamic data for arsenic were taken from data base of program MINTEQA2 (Allison et al. 1991). Multivariate statistics analysis was performed to verify the hydrogeochemical similarities among geochemical parameters. The data were analyzed using multivariate statistics implemented in the program PAST (Hammer et al. 2001). 4

GENERAL HYDROGEOCHEMISTRY

Selected results of the groundwater chemical analyses are presented in Table 2. In addition, Bhattacharya et al. 2004 (in press) provide more detailed discussion on the trends of spatial variability of water chemistry. Shallow groundwater (⬍50 m) in Brahmanbaria region was of Ca-Mg-HCO3 and Ca-Na-HCO3 types (Fig. 3a). Groundwater samples had very variable HCO⫺ 3 (74–562 mg/L) and SO 42⫺ (bdl-32.9 mg/L) concentrations. In the intermediate and deeper aquifers groundwater of Na-Cl-HCO3 type was also found (Fig. 3b). Groundwater pH values were between 6.2 and 7.6. Field Eh values corrected with respect to hydrogen electrode were from ⫹0.180 to ⫹0.29 V, indicating moderately reducing conditions. However, these results do not represent redox status of groundwater, possibly because of aeration of groundwater in hand-pump wells and during pumping as discussed later. Concentrations of total arsenic (Astot) in shallow wells varied from 10 to 335 ␮g/L and in intermediate wells reached up to 439 ␮g/L. Concentrations of dissolved Fe were highly variable, from 0.28 mg/L to 10.3 mg/L. However, no correlation between dissolved iron (Fetot) and Astot in shallow samples and only low correlation in intermediate depth samples were observed (Bhattacharya et al. 2004, in press). Most of dissolved arsenic (up to 99.5%) was present as As(III). Concentrations 136

137

Figure 3. Major ion characteristics of Brahmanbaria groundwaters plotted on a piper diagram . (a) Shallow wells (⬍50 ) (b) Intermediate wells (50–150 m, black circles) and deep wells (⬎150 m, data not included in discussion).

Figure 4.

Spatial variability of total As (Astot) in the shallow wells of Brahmanbaria in eastern Bangladesh.

of NH⫹ 4 were high in some shallow wells, reaching 12.2 mg/L. High concentrations of dissolved organic matter (DOC, up to 21.8 mg/L) were consistent with reducing character of groundwater. Dissolved sulfide with concentrations up to 2.1 mg/L was found in several wells, indicating the presence of sulfate reduction. Spatial variability of the distribution of arsenic in the shallow groundwaters and its relationship with other chemical parameters is discussed in detail in Bhattacharya et al. (2004, in press). Three domains were defined: Domain 1 with high concentrations Astot, and PO 43⫺ and low Fetot, and anomalous SO 42⫺ concentrations; Domain 2 with high concentrations of Astot, and PO 43⫺, and low concentrations of Fetot, and sulfate; Domain 3 with low concentrations of Astot, and PO4, and with high Fetot, and sulfate concentrations (Fig. 4). 5

GEOCHEMICAL MODELLING

Results of calculations of saturation indices (SI) together with calculated log PCO2 values are presented in Table 3. There was no significant complexation of Fe with other inorganic anions and the 138

Table 3.

Results of speciation calculations.

Sample ID

Depth (m)

SIsiderite

SIvivianite

SI rhodochrosite

logPCO2 (atm)

DIC (mol/L)

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30

18.3 21.3 21.3 21.3 21.3 22.9 24.4 24.4 25.9 27.4 27.4 27.4 27.4 28.0 29.0 29.0 32.0 32.0 32.0 32.0 39.6 41.1 54.9 59.4 62.5 64.0 64.0 68.6 73.2 75.6

⫺0.66 0.81 ⫺0.52 0.33 ⫺0.02 ⫺1.13 0.47 0.15 ⫺0.57 0.92 0.05 0.72 0.11 ⫺0.22 0.59 0.16 ⫺0.07 0.39 0.34 0.05 0.68 ⫺0.46 0.94 0.08 ⫺0.55 0.82 0.12 ⫺0.73 ⫺0.10 ⫺0.15

⫺3.80 2.57 ⫺2.31 0.46 ⫺0.72 ⫺4.29 ⫺0.14 ⫺1.23 ⫺2.72 2.36 ⫺1.65 ⫺0.70 ⫺0.38 ⫺1.56 2.45 ⫺1.11 0.01 ⫺0.43 0.40 ⫺1.06 2.04 ⫺1.92 1.80 ⫺0.03 ⫺1.91 2.66 ⫺0.61 ⫺4.52 ⫺1.86 ⫺0.73

⫺0.50 0.57 ⫺0.12 ⫺1.11 ⫺1.48 ⫺1.79 ⫺0.70 ⫺1.51 ⫺1.78 0.17 ⫺1.14 ⫺0.73 ⫺1.42 ⫺1.68 0.51 ⫺1.15 ⫺0.47 ⫺0.99 ⫺0.69 ⫺1.44 ⫺0.52 ⫺1.56 ⫺0.68 ⫺1.51 ⫺1.41 0.48 ⫺1.62 ⫺1.10 ⫺2.12 ⫺1.49

⫺1.94 ⫺2.37 ⫺2.41 ⫺1.10 ⫺0.90 ⫺1.63 ⫺1.20 ⫺1.19 ⫺0.93 ⫺2.09 ⫺1.67 ⫺1.47 ⫺1.30 ⫺0.82 ⫺2.78 ⫺1.00 ⫺1.96 ⫺1.48 ⫺1.37 ⫺0.78 ⫺1.68 ⫺1.13 ⫺1.48 ⫺1.31 ⫺1.42 ⫺2.18 ⫺0.98 ⫺1.17 ⫺1.05 ⫺1.28

2.73 ⫻ 10⫺3 2.37 ⫻ 10⫺3 2.09 ⫻ 10⫺3 1.10 ⫻ 10⫺2 1.09 ⫻ 10⫺2 2.31 ⫻ 10⫺3 8.73 ⫻ 10⫺3 4.82 ⫻ 10⫺3 1.01 ⫻ 10⫺2 3.62 ⫻ 10⫺3 1.01 ⫻ 10⫺2 3.96 ⫻ 10⫺3 6.90 ⫻ 10⫺3 1.14 ⫻ 10⫺2 1.16 ⫻ 10⫺3 1.01 ⫻ 10⫺2 2.04 ⫻ 10⫺3 8.23 ⫻ 10⫺3 8.66 ⫻ 10⫺3 1.24 ⫻ 10⫺2 9.74 ⫻ 10⫺3 1.03 ⫻ 10⫺2 8.26 ⫻ 10⫺3 1.24 ⫻ 10⫺2 1.13 ⫻ 10⫺2 2.45 ⫻ 10⫺3 1.27 ⫻ 10⫺2 8.01 ⫻ 10⫺3 1.27 ⫻ 10⫺2 8.61 ⫻ 10⫺3

principal aqueous species of Fe was Fe2⫹ and minor species was FeHCO⫹ 3 . Concentrations of Fe(III) were low, and the principal species were Fe(OH) 03 and, at lower pH, Fe(OH)⫹2 . Low Mn concentrations in groundwater within the reduced domains could possibly be due to precipitation of rhodochrosite (MnCO3) (Sracek et al. 2000, McArthur et al. 2001, Ahmed et al. 2004). Calculated log PCO2 values were very high, reaching in some cases values higher than – 1.0. This is related to the generation of CO2 in redox reactions like dissolution of ferric oxide and hydroxides in reaction with organic matter. This is consistent with high calculated values of DIC (up to 1.27 ⫻ 10⫺2 mol/L). Many samples (60%) were at equilibrium or supersaturated with respect to siderite (FeCO3) suggesting that this mineral phase might have acted as a sink for dissolved iron. Some samples (30%) were also supersaturated with respect to vivianite Fe3(PO4)2.8H2O, but the degree of saturation is generally lower than in the case of siderite. Few samples (13.3%) are also supersaturated with respect to rhodochrosite. Some samples are close to equilibrium with calcite and dolomite (not shown), but saturation is reached only in very limited number of samples. The speciation program was also used to calculate Eh values based on As(III)/As(V) couple and S(VI)/S(-II) couple determined analytically. Results of these calculations are presented in Table 4. Typical feature is the strong disequilibrium between measured field Eh values adjusted with respect to hydrogen electrode and values of Eh calculated on the basis of arsenic couple. Values of redox potential based on arsenate to arsenite ratios are lower than the field Eh values. This holds even more for sulfur redox couple, suggesting strong redox disequilibrium. However, the field Eh values truncated at ⫹0.180 V are possibly unreliable and must be interpreted with caution. There 139

Table 4.

Comparison of redox data.

pH

SO42-

Mg2+

Ca2+

EC

3.09 1.66 1.96 2.14 1.55

K+

⫺0.252 ⫺0.205 ⫺0.225 ⫺0.247 ⬍⫺0.254

PO43-

0.029 0.059 ⫺0.011 ⫺0.073 ⫺0.018

Fetot

⫹0.180 ⫹0.180 ⫹0.180 ⫹0.180 ⫹0.180

Mn

3 16 19 21 24

Astot

SImackinawite

NO-3

Eh (V) S(VI)/S(-II)

Na+

Eh (V) As(V)/As(III)

CI-

Field Eh (V)

HCO3-

Sample ID/ Parameter

1

2

3

4

5

6

7

8

9

10

11

12

13

14

0 -100 -200

Similarity

-300 -400 -500 -600 -700 -800

Figure 5.

Results of Hierarchical Cluster Analysis (HCA).

also is a possibility of precipitation of secondary sulfide minerals like mackinawite. This mineral is a precursor of pyrite and may incorporate some arsenic, acting as a sink for arsenic in groundwater where sulfate reduction takes place.

6

MULTIVARIATE STATISTICS

Results of Hierarchical Cluster Analysis (HCA) performed in Ward’s mode using the program PAST are given in Figure 5. They indicate highest degree of similarity among NO⫺ 3 , Mn, Fetot, PO 43⫺, SO 42⫺, Astot, and pH. Most of them are redox sensitive species, except for PO 43⫺ which is linked to Fe due to its release during reductive dissolution of ferric oxide and hydroxides. The effect of pH is however not clear, although this parameter plays a role in precipitation of minerals like siderite and vivianite. Na⫹ and Cl⫺ form a distinct group, which most likely indicates relict sea water entrapped in the aquifers. Ca2⫹ and Mg2⫹ are separated from Na⫹ and Cl⫺ because they are related not only to relict sea water in the sediments, but also to processes like dissolution of carbonates, and weathering of silicates enhanced by production of CO2 in redox reactions. HCO⫺ 3 140

400 11 25 4 22

300 9

21

18 20 19 16 23 7 5 14

24

13

200

29

Component 2

30 10 8 12

100

27 17

3 6

0

1 2 26

15

-100

-200 28

2000

1000

3000

Component 1

Figure 6.

Results of Principal Components Analysis (PCA).

generated in several redox reactions and carbonate dissolution is linked to almost all parameters and this holds even more for EC. Results of Principal Components Analysis (PCA) presented in Figure 6 indicate two principal components and identified as: PC1 with high loadings for EC, Na⫹, Cl⫺, and HCO ⫺ 3 , and PC2 and with relatively low, but significant loading for arsenic. The PC1 with high loading for HCO ⫺ 3 corresponds to the influence of relict seawater entrapped in the sediments to the groundwater, and PC2, which corresponds to the impact of redox reactions. These principal components explain 92.01% and 6.71%, respectively, of total variance in sample set. The samples at the bottom right (27 and 28, Fig. 6) indicate strong influence of saline water, most likely reflect to be relict seawater entrapped in the sediments. On the other hand, samples at the top of the graph (4, 11, 21, 22, 25 etc., Fig. 6) are strongly influenced by redox reactions and they generally have high arsenic concentrations. Samples at the bottom left (1, 2, 3, 6, 17 etc., Fig. 6) are relatively less influenced by both processes and have low arsenic concentrations. It seems that redox processes relatively less influence the samples showing signature of relict saline water from the marine sources. However, the impact of palaeo-seawater relicts in the aquifers seems to be a local phenomenon in the Brahmanbaria area, which is also seen in many other areas of Bangladesh.

7

DISCUSSION AND CONCLUSIONS

The conceptual model of arsenic and iron behavior in groundwater in Bangladesh can be summarized as follows: (a) reductive dissolution of ferric oxide and hydroxides in reaction with organic matter like peat after consumption of more favored electron acceptors in a reaction like:

141

where CH2O represents simplified organic matter. This reaction produces dissolved Fe(II), increases DIC concentration, and raises pH. Also, arsenic initially present on adsorption sites is released to groundwater. (b) Precipitation of Fe(II) as siderite, FeCO3, and vivianite, Fe3(PO4)2.8H2O. These processes decrease dissolved Fe(II) and, in the case of siderite precipitation, there is reduction of DIC concentration. However, the impact on DIC and bicarbonate concentrations is limited because their pools are much larger than dissolved iron pool. When sulfate reduction takes place, minerals like mackinawite also act as active sinks for dissolved iron. However, sulfate concentrations are generally low and this process does not seem to be very significant. This means that correlation between dissolved iron and arsenic frequently observed at sites contaminated by arsenic may be disturbed because behavior of dissolved arsenic is more conservative than behavior of dissolved iron. There are no important minerals of arsenic in groundwater with low sulfate concentrations and adsorption of arsenic is limited in alkaline pH region (Langmuir 1997). Results of speciation modeling are consistent with the conceptual model (Bhattacharya et al. 1997, Nickson et al. 2000, Bhattacharya et al. 2002a,b, Ahmed et al. 2004) because there are very high PCO2 and DIC values in samples with high arsenic concentrations. Also, many samples are supersaturated with respect to siderite, and, to less extent with respect to vivianite. Presence of ferrous carbonate and phosphate minerals was also confirmed by sequential extraction (Ahmed et al. 2004). Multivariate statistics groups together redox-sensitive parameters like Fe, Mn, As, and DOC. Phosphate is also in the same group because it is indirectly linked to redox processes through its release in reductive dissolution of ferric oxide and hydroxides. Potential initial source of phosphate is decomposition of peat as well as organic matter disseminated in the alluvial sediments (Routh et al. 2000, Ravenscroft et al. 2001). In contrast, there is less resemblance with bicarbonate, which can also be produced by other processes like dissolution of soil calcite. Multivariate statistics also helped to identify samples with a significant impact of sea water. Samples with a significant seawater impact had high loadings for typical sea water ions such as Na⫹ and Cl⫺. Sulfate was less useful indicator because its concentration was very low. The combination of geochemical modeling and multivariate statistics has been proved to be a useful tool in testing the hypothesis about arsenic release mechanism.

ACKNOWLEDGEMENTS The authors would like to acknowledge the Swedish Research Council (VR) and Sida-SAREC for supporting the research on High arsenic groundwater in Bangladesh since January 1997. The authors would like to thank Andreas Mende for his valuable suggestions to improve the manuscript.

REFERENCES Ahmed, K.M., Hoque, M., Hasan, M.K., Ravenscroft, P. & Chowdhury, L.R. 1998. Origin and occurrence of water well methane gas in Bangladesh aquifers. Jour. Geol. Soc. India 51: 697–708. Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A., Imam, M.B., Khan, A.A & Sracek, O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh: An overview. In P Bhattacharya, AH Welch, KM Ahmed, G Jacks, R Naidu (eds.) Special Issue: Arsenic in Groundwater of Sedimentary Aquifers, Appl. Geochem. 19(2): 181–200. Allison, J.D., Brown, D.S. & Novo-Gradac, K.J. 1991. MINTEQA2, A Geochemical Assessment Data Base and Test Cases for Environmental Systems, Athens, GA, U.S. EPA. Bakr, M.A. 1977. Quaternary Geomorphic Evolution of the Brahmanbaria-Noakhali Area, Comillaand Noakhali Distrcits, Bnagladesh. Records of the Geological Survey of Bnagladesh, Volume 1, part 2, 48p. BGS & DPHE 2001. Arsenic Contamination of Groundwater in Bangladesh. Vol 2 Final Report, BGS Technical Report WC/00/19.

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Bhattacharya, P., Chatterjee, D. & Jacks, G. 1997., Occurrence of arsenic contaminated groundwater in alluvial aquifers from Delta Plains, Eastern India: Options for safe drinking water supply. Int Jour. Water Res. Management 13(1): 79–82. Bhattacharya, P., Frisbie, S.H., Smith, E., Naidu, R., Jacks, G. & Sarkar B. 2002a. Arsenic in the Environment: A Global Perspective. In B.Sarkar (ed.) Handbook of Heavy Metals in the Environment (Chapter 6), New York: Marcell Dekker Inc., pp. 145–215. Bhattacharya, P., Jacks, G., Ahmed, K.M., Khan, A.A. & Routh, J. 2002b. Arsenic in groundwater of the Bengal Delta Plain aquifers in Bangladesh. Bull. Env. Cont. Toxicology 69: 538–545. Bhattacharya, P., Ahmed, K.M., Broms, S., Fogelström, J., Jacks, G., Sracek, O. & Routh, J. 2004. Mobility of arsenic in groundwater in a part of Brahmanbaria district, NE Bangladesh In: Naidu, R., Smith, E., Smith, L., Smith, J. and Bhattacharya, P. (Eds.) Managing Arsenic in the Environment: From soil to human health. CSIRO Publishing, Melbourne, Australia. (In press) Brammer, H. 1996. The Geography of the Soils of Bangladesh. University Press Ltd., Dhaka, Bangladesh. Hammer, Ø., Harper, D.A.T. & Ryan, P.D. 2001. Past: Paleontological Statistics Software Package for Educational and Data Analysis. Paleontologica Electronica 4(1): 9 p. Langmuir, D. 1997. Aqueous Environmental Geochemistry, Prentice Hall, New Jersey. McArthur, J.M., Ravencroft, P., Safiullah, S. & Thirlwall, M.F. 2001. Arsenic in groundwater: testing pollution mechanism for sedimentary aquifers in Bangladesh. Water Resour. Res. 37: 109–117. Meng, X., Korfiatis, G.P., Christodoulatos, C. & Bang, S. 2001. Treatment of arsenic in Bangladesh well water using a household co-precipitation and filtration system. Wat. Res. 35: 2805–2810. Mukherjee, A.B. & Bhattacharya, P. 2001. Arsenic in groundwater in the Bengal Delta Plain: Slow Poisoning in Bangladesh. Environmental Reviews 9(3): 189–220. Nickson, RT, McArthur, J.M., Ravenscroft, P., Burgess, W.G. & Ahmed, K.M. 2000. Mechanism of arsenic release to groundwater, Bangladesh and West Bengal. Appl. Geochem. 15(4): 403–413. Parkhurst, D.L. 1995. Users Guide to PHREEQC-A Computer Program for Speciation, Reaction-Path, Advective-Transport, and Inverse Geochemical Calculations, U.S. Geological Survey Water Resources Investigation Report 95–4227. Ravenscroft, P., McArthur, J.M. & Hoque, B.A. 2001. Geochemical and paleohydrological controls on pollution of groundwater by arsenic. In Chappell, W.R., Abernathy, C.O. & Calderon R.L. (eds.) Fourth International Conference on Arsenic Exposure and Health Effects. Routh, J., Bhattacharya, P., Jacks, G., Ahmed, K.M., Khan, A.A. & Rahman, M.M. 2000. Arsenic geochemistry of Tala groundwater and sediments from Satkhira District, Bangladesh. Eos Trans American Geophysical Union 81(48): 550 Smedley, P. & Kinniburgh, D.G. 2002. A review of the source, behavior and distribution of arsenic in natural waters. Appl. Geochem. 17: 517–568. Sracek, A., Bhattacharya, P., Jacks, G., Chaterjee, D., Larsson, M. & Liess, A. 2000. Groundwater arsenic in the Bengal delta Plains: A sedimentary geochemical overview. In: A.L. Ramanathan, V. Subramanian & R. Ramesh (eds.) Proc. International Seminar on Applied Hydrogeochemistry’ Annamalan University, Tamil Nadu, India, pp. 47–56. Umitsu, M. 1987. Late Quaternary sedimentary environment and landform evolution in the Bengal Lowland. Geog. Rev. Japan (Ser. B) 60: 164–178. Umitsu, M. 1993. Late Quaternary sedimentary environments and landforms in the Ganges Delta. Sed. Geol. 83, 177–186 van Geen, A., Ahmed, K.M., Seddique, A.A. & Shamsudduha, M. 2003. Community wells to mitigate arsenic crisis in Bangladesh. Bulletin of the World Health Organization 81(9): 632–638. WHO 2001. Arsenic in drinking water. Fact sheet 210: URL: http://www.who.int/mediacentre/factsheets/fs210/en/print.html (Accessed on March 9, 2004)

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Microbial processes and arsenic mobilization in mine tailings and shallow aquifers J. Routh Department of Geology and Geochemistry, Stockholm University, Stockholm, Sweden

A. Saraswathy Department of Biology, West Virginia State College, West Virginia, USA

ABSTRACT: Microbial processes play an important role in transforming and mobilizing As in the sub-surface. Enrichment cultures indicated several As tolerant species, which actively reduced As(V) to As(III). No change in As speciation occurred in the controls, thereby confirming As(V) reduction was biologically mediated, and active metabolism was a prerequisite for reduction. Different growth and As(V) reduction rates were noted under oxic to sub-oxic conditions, and a zero-order model best fits the As(V) reduction data. Arsenic concentrations in the microcosms seem to affect biomass yield and As(V) reduction rates in some of the strains. Arsenic reduction in these microorganisms probably occurs for respiratory or detoxification purposes. It is likely that microbial mobilization of As may have an impact on groundwater remediation treatment in these environments.

1

INTRODUCTION

Arsenic (As) compounds are highly toxic (Nriagu 2002). People have known about its lethal properties since antiquity, and have often used arsenic trioxide (As2O3) for homicide purposes. Renewed interest in the metalloid has increased dramatically, but for different reasons. A recent paper by Bhattacharya et al. (2002) details the As crisis, which affects the lives of general population en masse in several countries. The problem is nowhere as serious as it is in Bangladesh and India, where more than 70 million people are at risk from drinking As-rich groundwater (Harvey et al. 2003). The first reports on clinical manifestation of As toxicity from this region came up around 1978, and thereafter, chronic cases of As poisoning (including fatalities) have been reported since 1982 (Saha 1984, Goriar 1984). Arsenic commonly occurs in the environment as inorganic trivalent As(III) and pentavalent As(V) species (Cullen & Reimer 1989). The trivalent arsenous acid is more dominant under reducing conditions, whereas its pentavalent counterpart, in the form of arsenic acid is common under oxidizing conditions. Mobility of As(III) is higher compared to As(V) in sedimentary and aqueous environments. The difference in mobility is attributed to the high affinity of As(V) for insoluble species such as hydrous ferric and manganese oxides (Cullen & Reimer 1989). Additionally, several methylated forms of organoarsenicals (e.g., methylarsonic, methylarsonus, and dimethylarsenic acid) are also found in water as breakdown or excretory byproducts (Cullen & Reimer 1989, Sohrin et al. 1997). Although As(III) is not thermodynamically stable under oxidizing conditions, there are several incidences where As(III) occurs as the dominant species in the water column (Aurillo et al. 1994, Sohrin et al. 1997). Prevalence of As(III) in these studies was correlated with phytoplankton abundance suggesting non-equilibrium conditions were microbially mediated. Arsenic enters the terrestrial and aquatic environments through both natural and anthropogenic activities. Natural processes can contribute to the widespread distribution of As through weathering 145

of As bearing rocks and minerals, microbial activity, and volcanic eruptions. In contrast, anthropogenic point sources are localized and include inputs from mining of base metals, smelter slag, coal combustion, production of paints and dyes, tanning, wood preservation, and pesticides. The net output of As from anthropogenic processes is high compared to natural processes (Bhattacharya et al. 2002), but ironically, it is the natural processes that are involved in dispersion of As, which is of most concern to human beings. 1.1

Arsenic toxicity and microbial resistance

Arsenic compounds readily accumulate in living tissues due to their affinity for proteins and lipids or cause breakdown of oxidative phosphorylation (Oremland & Stolz 2003). Many prokaryotes and eukaryotes have however, developed unique inter-cellular reaction mechanisms to rid themselves of As, and excrete it as waste byproducts. Some even generate energy during this process. For example, higher eukaryotes reduce As(V) to As(III) followed by methylation resulting in mono and dimethylarsonoic acids (MMA, DMA). Fungi produce trimethylarsines and bacteria produce MMA and DMA (Diorio et al. 1995, Sohrin et al. 1997). The physical excretion of these byproducts often occur as encrustations on the cell wall (Saraswathy et al. 2004) or transferred into the water column at the sediment-water interface (Ahmann et al. 1997, Martin & Pedersen, 2002, Routh et al. 2004). Additionally, As can be converted to benign products as arsenobetaine and As containing sugars found in marine algae, animals, and higher plants (Cullen & Reimer 1989). Reduction of As(V) to As(III) in anoxic environments primarily occurs via dissimilatory As reduction (Ahmann et al. 1997, Newman et al. 1997a,b), whereby microorganisms utilize As(V) as the terminal electron acceptor. The reaction is energetically favorable and coupled to oxidation of organic matter. Two important prerequisites for such microorganisms are the presence of strict anoxic conditions and high As levels (Newman et al. 1997b). Dissimilatory As reduction occurs in bacteria scattered throughout the bacterial domain representing ␥-, ␦-, ␧-Proteobacteria, low-GC gram-positive bacteria, thermophilic Eubacteria, and Crenoarchea (Oremland & Stolz 2003). However, microorganisms may also possess reduction mechanisms that are not coupled to respiratory processes, but instead, impart resistance to As toxicity (e.g., Jones et al. 2000, Macur et al. 2001). Enzymes involved in the detoxification pathway are transcribed by the ars operon resulting in inter-cellular reduction of As(V), and the subsequent efflux of As(III) via a trans-membrane pump (Cervantes et al. 1994, Rosen 2002). 1.2

Arsenic mobilization

Different biogeochemical mechanisms have been proposed to explain As mobilization in sedimentary environments. These processes involve: (1) oxidation of pyrite, (2) reduction of Fe-oxyhydroxides coupled to oxidation of organic matter and release of As(V), (3) exchange of As(V) with phosphate based fertilizers (e.g., Roy Chowdury et al. 1999, Nickson et al. 1998, Harvey et al. 2003). More recently, researchers have increasingly focused on the role of microorganisms in mobilizing As in sedimentary environments (e.g., Ahmann et al. 1997, Cummings et al. 1999, Jones et al. 2000, Macur et al. 2001, Islam et al. 2004). Transformation of As by microorganisms has important environmental implications because As(V) and As(III) have different sorption and toxicological properties. Primarily such studies have mostly focused on sites contaminated by mining, pesticides or other related anthropogenic activities, and they all demonstrate enhanced microbial As mobilization on short time scales. Here, we present data from our ongoing investigations on mine tailings in northern Sweden and shallow aquifers in West Bengal, India. The environments are completely different in terms of physiographic settings, sub-surface geology, and environmental conditions, but both places indicate high As concentrations in ground and surface water. Although different processes in the sub-surface may substantially influence the biogeochemical cycling of As, we focused on microbial processes and their affect As cycling. The different microbial processes important in As cycling underscore the need for our continued inquiry regarding As transformations, in hopes, that we can develop 146

greater predictability of its behavior. To the best of our knowledge, microbial processes affecting As mobilization has not been investigated by other researchers at these sites, and may thus, provide new insights. Moreover, microbial processes involved in As(V) reduction and mobilization are many times faster than chemical transformation (e.g., Sohrin et al. 1997, Jones et al. 2000). If microbial processes are indeed active, this may have important environmental implications on As remediation for groundwater treatment and management at these sites.

2 2.1

MATERIALS AND METHODS Sampling sites

Adak mine tailings: The abandoned mine tailings at Adak in Västerbotten district of northern Sweden extends over 1500 m ⫻ 2000 m ⫻ 5 m. The site is affected by acid mine drainage and has low pH (⬃3–4). The tailings have high concentrations of As, Cu, and Zn, and they have been mixed with glacial till to reduce surficial weathering (Jacks et al. 2003). The tailings are underlain by peat bogs, and extend close to the shores of Lake Ruttjejaure. Shallow streams running adjacent to the tailings deposits, drain into Lake Ruttjejaure carrying washouts of mine tailings. Sampling was done using a gravity corer to extract undisturbed sediment cores from the lake. Details on sampling and different geochemical and microbiological assays are further discussed in Bhattacharya (2004). Ambikanagar groundwater aquifer: Ambikanagar is located in the Deganga Block of North 24 Parganas in West Bengal, India. The water table occurs at a depth of 5-m, and rainfall in summer is the principal source of recharge for aquifers. Reconnaissance work by our group indicated that As concentrations in the shallow wells are often above the permissible drinking water limit (50 ␮g/L; Routh et al. 2003). The underlying thick Quaternary alluvium consists of cycles of complete or partly truncated fining-upward sequences dominated by coarse to medium sand, fine sand, silt, and clay. Aquifer sediments in the deep wells are mostly coarse sands, whereas the shallow wells usually consist of fine-to-medium grained sands. Air jet drilling was used to install an 18-m deep well. The aquifer sediments were collected in Anero™ (Mitsubishi, Inc.) bags and shipped to Stockholm for microbial assays. The experiments were started within 4-days after sampling. 2.2

Microcosm experiments

The microcosm experiments were a two-step process: (1) enrichment studies to isolate bacteria tolerant to high As levels, and (2) determination of the As mobilizing capacity of isolated pure microbial cultures. The sediments were inoculated into sterile minimal medium (Turpeinen et al. 1999). The medium contained lactate as the carbon source, and it was spiked with As. The microcosms were sampled, and a specific volume of sample sacrificed periodically for different microbiological and geochemical assays. The sediment slurries extracted under aseptic conditions through the rubber septa were centrifuged at 907 G (3000 rpm). The aqueous and sediment phases obtained were separately analyzed for As(III) and As(V) species, and compared to the heterotrophic plate counts of the corresponding day. The heterotrophic plate counts were performed by serially diluting the samples using phosphate buffered saline solution. The bacterial culture was spread on Tryptic Soy Agar plates spiked with As, and incubated for 72 hrs at 22°C. The bacterial colonies were selected and replated until pure cultures were obtained. The isolates were identified using the API and 16S rRNA techniques. The pure cultures were inoculated into sterile basal salts medium under oxic to sub-oxic (2–3 mg/L dissolved oxygen; obtained by bubbling Ar through the medium and storing samples in N2 filled box) conditions containing 1 mM, 2 mM, and 5 mM As(V) and lactate. During sampling Eh, pH, and oxygen were measured in the microcosms. Replicate samples were sacrificed periodically. Microbial growth was determined by measuring change in optical density (600 nm) and dry weight over the duration of the microcosm experiment. As(III) and As(V) species in the medium 147

was measured by modifying existing spectrophotometric methods (Johnson & Pilson 1972, Cummings et al. 1999). Optical density was calibrated for each strain. Additional samples were set up as controls after treating the samples with HgCl2 and formaldehyde. Specific details of these procedures are provided in Collins et al. (2004) and Routh et al. (2004).

3

RESULTS

3.1 Adak mine tailings Microbes enhanced dissolution of As in enrichment cultures by increasing As(III) concentrations in the aqueous phase and mobilizing ⬃27–51% of As present in contaminated sediments (Bhattacharya et al. 2003). Several bacteria were isolated from the enrichment studies, and identification of different microbial strains is ongoing. Here, we have focused on two microbial strains where we have generated complete data for: (1) arsenic transformation and mobilization, and (2) API and 16S rRNA identification. Arsenicicoccus bolidensis is hitherto a new species of actinomycete and it is a gram-positive, facultatively anaerobic, coccus-shaped microorganism (Fig. 1; Collins et al. 2004). The microcosm experiments indicated a fall in As(V) coupled to increase in As(III) and heterotrophic growth (indicated as increase in optical density and dry weight). A. bolidensis reduced 0.06–0.20 mM/day As(V) under sub-oxic conditions (Saraswathy et al. 2004). Arsenic reduced by the bacteria occurs as encrustations on bacterial cells as shown by EDAX X-emission spectrum (Fig. 1). The As(V) reduction values are low compared to other As reducing microorganisms (Routh et al. 2004). As(V) reduction is however, related to growth in A. bolidensis implying that respiration and/or detoxification pathways may be involved in As(V) transformation. Notably, this is the first report of an actinomycete involved in As reduction in sedimentary environments. A novel species of facultatively anaerobic Chromobacterium occurring as rod-shaped microorganisms were isolated from the Adak sediments (Fig. 1). Bacterial growth in the culture decreased after 12 days corresponding to the conversion of 74% of 1 mM As(V) to As(III). This bacterium was able to reduce 0.7–0.22 mM/day of As(V) (Fig. 2). Oxygen levels as high as 0.025 mM did not affect bacterial growth or As(V) reduction. In the presence of other electron acceptors in competition experiments involving lactate and a combination of As(V) with sulfate or nitrate, only As(V) concentrations varied. 3.2 Ambikanagar shallow aquifer Enrichment cultures indicated several As tolerant species, which actively reduced As(V). Specific details regarding the geochemical trends indicated by these bacteria are discussed in Routh et al. (2004). Continued increase in As concentrations in the enrichment cultures affected bacterial growth resulting in a decrease in plate counts and As(V) reduction. During the experiment, oxygen and Eh levels decreased, whereas pH increased. Amongst the eleven microbial strains isolated from the TSA plates, five of them were morphologically most distinct. These strains were selected for 16S rRNA characterization and As mobilization experiments. The bacteria were identified as: Acinetobacter johnsonii, Citrobacter freundii, Comamonas testosteroni, Enterobacter cloacae, and Sphingobium yanoikuyae. These bacteria range from aerobic to facultative anaerobic species. Similarity of the 16S rRNA sequence with GenBank varies between 98 and 100%. Different growth and As(V) reduction rates were noted on inoculating the basal salts medium. Maximum growth and As(V) reduction was noted in the bacteria A. johnsonii (Fig. 2), which corresponded with initial As(V) concentration and biomass yield. Compared to other As(V) reducing microorganisms, bacteria isolated in this study indicated lower reduction rates (0.11–0.25 mM/day). Notably, C. testosteroni and S. yanoikuyae did not indicate a direct correlation between As(V) concentration versus growth rate and biomass yield. It is likely that As(V) reduction in these bacteria was related to detoxification (e.g. Macur et al. 2001). 148

Figure 1. ESEM imaging in wet mode illustrating the Arsenicicoccus bolidensis and Chromobacterium cluster. Representative EDAX X-emission spectrum quantification collected from electron dense particles on bacterial surface as encrustations (inset). The S, O, and P speaks are due to background from the supporting grid. Chromobacterium

0.6

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1.0 As(V) As(III) O.D

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O.D. (600 nm)

4

0.8

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Acinetobacterjohnsonii 6

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As(V) As(III) O.D

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Figure 2. Growth of Chromobacterium and Acinetobacter johnsonii in minimal medium with 5 mM As(V). Note that change in As speciation corresponds with enhanced growth represented as optical density measurements in the microcosm cultures (experiments were conducted in duplicate).

4 4.1

DISCUSSION As(V) reduction and growth

Our investigations on As cycling in the mine tailings and shallow aquifers show several interesting results. First, in situ microorganisms play an important role in transforming and mobilizing As at both locations. The microbial strains indicated general resistance to As toxicity under oxic to sub-oxic conditions, and reduced As(V) to As(III). Increase in optical density and dry weight was directly correlated to growth in the microcosms inoculated with individual microbial strains. Notably, As speciation remained unchanged and no microbial growth occurred in the controls (data not shown). This clearly confirms that As(V) reduction was ‘biologically mediated’ and microorganisms play a role in As cycling. Because the enrichment cultures did not isolate iron and sulfate reducing bacteria (which are also capable of mobilizing As e.g., Cummings et al. 2000, Jones et al. 2000, Kuhn & Sigg 1993) their role in As cycling at these sites can only be speculative. Although we have clearly established As(V) reduction and mobilization in the sediment microcosms in absence of Fe and sulfur reducing bacteria, it is unknown if this process is affective in natural environments. While some of these microorganisms isolated in this study are new (e.g., 149

A. bolidensis and Chromobacterium), others have been associated with As cycling in sedimentary environments (e.g. S. yanoikuyae; Macur et al. 2001). Inasmuch as it is important to expect other biogeochemical conditions to play a role in affecting As(V) reduction, in situ microbial processes are probably more affective on short-term basis. Researchers have indicated that abiotic reduction of As(V) may occur due to sulfides (Kuhn & Sigg 1993), but no odor for H2S was detected during sampling or change in color noted in the sediments to suggest presence of iron sulfides. Moreover, both laboratory and field measurements indicate that abiotic reduction of As(V) is kinetically a slow process (Kuhn & Sigg 1993, Newman et al. 1997b), and does not match with the As transformation rates in these sediments (e.g. Routh et al. 2004). The effect of As(V) concentration on bacterial growth was evaluated. A zero-order model with respect to As(V) concentration best fits our experimental data. The kinetic model applied to evaluate As(V) reduction is similar to U(VI) reduction involving sulfate reducing bacteria (Spear et al. 2000). The model is represented as: (1) where As ⫽ is the model predicted As(V) concentration, As0 ⫽ initial concentration of As(V), k0 ⫽ maximum specific reaction rate coefficient expressed as As(V) concentration/mg (dry weight) of cells/ml/h, X ⫽ bacterial cell concentration in mg (dry weight)/ml, and t ⫽ time in hours. Figure 3 indicates the trend for modeled and fitted k0 values for Chromobacterium in microcosms containing 1, 2, and 5 mM As(V). The zero-order model is a simplification of Michaelis-Menten and Monod type kinetics at high substrate concentrations denoted by: (2) where Vmax ⫽ Michaelis-Menten maximum substrate utilization rate constant expressed as the As(V) concentration/mg (dry weight) of cells/ml/day, ␮m ⫽ Monod maximum specific growth rate constant/h, and Y ⫽ cell yield expressed as mass of cells in mg per mg of substrate used. 6 Observed 1 mM Fitted 1 mM

As(V) reduction (mM)

5

Observed 2 mM Fitted 2 mM

4

Observed 5 mM Fitted 5 mM

3 2 1 0 0

3

6

9

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15

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Figure 3. Time course of As(V) reduction by Chromobacterium first with a zero-order model. The plot is based on calculation of As concentration and k0 values (see equations 1 and 2 in text). Each set of data points represents at least two experiments conducted over a 15-days period.

150

Microbial growth and reduction rates varied between individual species, but compared to other microorganisms the reduction rates were low (Routh et al. 2004). In this context, an important distinction from other laboratory simulated studies (e.g., Ahmann et al. 1997, Newman et al. 1997 a, b among others) is the fact that we used carbon levels (external inputs to the cultures) that was substantially low (0.01 mM versus 1 to 10 mM in other studies). While the low organic carbon concentration is more reflective of the natural organic matter content in these environments, it may have resulted in low As(V) reduction rates. Organic matter breakdown products in the microcosms were not measured, but consistent increase in pH, As(III), and microbial numbers support reaction pathways involving breakdown of lactate (e.g. Zobrist et al. 2000) or natural sedimentary organic matter. Other researchers have also indicated fall in As(V) reduction with decrease in organic substrate in microcosms (e.g. Ahmann et al. 1997, Harvey et al. 2003, Islam et al. 2004). The results reiterate the importance of organic matter for the survival of these heterotrophic microbial communities in the sub-surface. One of the confounding issues in this study is underpinning whether these microorganisms reduce As(V) for detoxification or respiratory purposes. This partly arises due to the aerobic or strict anaerobic habitats preferred by the respective microbial colonies involved in As(V) reduction (Newman et al. 1997b). The general conditions in this study were sub-oxic, and some of the microorganisms probably use As(V) as an alternate electron acceptor. Interestingly, recent studies by other researchers imply that presence of an enzymatic detoxification pathway does not preclude the As(V) respiration capability in bacteria. For example, both detoxifying and respiratory As (V) reductases occur in the microaerophile Bacillus selenitireducens (Switzer-Blum et al. 1998) and the As(V)-respiring anaerobe Shewanella ANA-3 strain (Saltikov & Newman 2003). Similar possibilities may exist in some of the bacteria discussed here (e.g., A. johnsonii, A. bolidensis), but genetic evidence (on same lines as in Saltikov & Newman 2003) is presently unavailable to support this idea. Nonetheless, the fact that some these microorganisms are capable of using O2 or switch to other terminal electron acceptor during respiration implies that they are generally opportunistic by character. By means of complex inter-cellular processes these microorganism are able to survive under conditions that are generally less preferable to others in the sub-surface. The study implies the potential impact such microorganisms could have on As cycling at these sites. 4.2

Environmental implications

Of late, the thrust by different regulatory agencies is on developing As remediation methods that are supposed to be cost-affective and reaches out to a larger population. The most commonly used in situ techniques involve maintaining aerobic conditions through aeration or using chemical oxidants to convert As(III) to the less mobile and toxic As(V) species. In this context, microbial transformation and mobilization of As in the sub-surface may have important implications on groundwater treatment. First, microbial As(V) reduction if it is common, then the prediction of As valence, and thus behavior, based solely on redox status may be problematic. For example, even under oxidizing conditions As(III) has been found as the predominant species in the Adak tailings deposit (Bhattacharya et al. 2003) similar to other studies (Aurillo et al. 1994, Sohrin et al. 1997). Second, efforts made to chemically oxidize As(III) to As(V) during groundwater treatment may be unproductive since As(V) is actively reduced to As(III) by in situ bacteria. This is largely bad news for researchers focusing on developing As remediation techniques. Most methods up to date have focused on manipulating the redox states and converting As(III) to As(V) (for review see Murcott 2001, Ahmed 2001). Many of these methods hardly take into account the role of in situ microbial activity. Hence, it is not surprising many groundwater treatment methods fail to work effectively under field conditions. Clearly, there is an urgent need to assess the occurrence and efficiency of these microbial processes in field-based pilot studies as a prerequisite to provide critical information before implementing specific remediation methods for removing As. 151

5

CONCLUSIONS

Microbial processes play a crucial role in As mobilization. Mobilization of As from sediments into the aqueous phase is mediated by eukaryotes, fungi, and bacteria. This involves reduction of As(V) into the more mobile and toxic As(III) species. Microbial reduction of As(V) mainly arises for detoxification or respiratory purposes. They are different biochemical pathways occurring under mostly oxic or strict anoxic conditions, respectively. Here, we have indicated the role of microorganisms in mobilizing As at sites contaminated by mine tailings (in northern Sweden) and shallow aquifers in West Bengal (India). We isolated several microorganisms (including two new species) that are involved in As reduction under oxic to suboxic conditions. These microorganisms are generally resistant to As toxicity. As (V) reduction in the microcosms, correspond with increase in dry weight and optical density measurements. In some of these microorganisms, the correspondence between As concentration, bacterial growth, and biomass yield is high. This leads us to believe that some of these microorganisms are probably using As(V) for respiration in addition to detoxification purposes. Further genetic work is required to understand such complex biochemical pathways. Nonetheless, the study proves that the microorganisms whether they are present in mine tailings or groundwater aquifers; they are generally opportunistic by character. The microorganisms have developed unique survival skills, and in the process, created a microbial niche for themselves. Enhanced mobilization of As from sediments has important implications on groundwater treatment. Active microbial processes may result in disequilibrium conditions, whereby even under oxidizing conditions As(III) species may predominate. Additionally, groundwater treatment based on aeration and oxidation processes (if implemented at these sites) needs to be assessed critically. Given the fact that in both places, we have a thriving microbial community in the sub-surface, which is able to reduce As(V) under oxic to sub-oxic conditions, it raises questions about groundwater treatment methods suitable for these sites. ACKNOWLEDGEMENT We thank the Geological Survey of Sweden (SGU) and Swedish International Development Agency (Sida-SAREC) for providing the research funds to conduct our studies in Sweden and India. Gunnar Jacks, Sisir Nag, S.P. Sinha Ray, and Prosun Bhattacharya helped us with fieldwork. We thank Roger Herbert and Jim Saunders for their suggestions. Rolf Hallberg helped with the ESEM imaging. REFERENCES Ahmann, D., Krumholz, L.R., Hemond, H.F., Lovley, D.R. & Morel, F.M.M. 1997. Microbial mobilization of arsenic from sediments of the Aberjona watershed. Environ. Sci. Tech. 31: 2923–2930. Aurillo, A.C., Mason, R.P. & Hemond, H.F. 1994. Speciation and fate of arsenic in three lakes of the Aberjona watershed. Environ. Sci. Tech. 28: 577–585. Bhattacharya, P., Jacks, G., Frisbie, S.H., Smith, E., Naidu, R. & Sarkar, B. 2002. Arsenic in the Environment: A Global Perspective. In B. Sarkar (ed.) Heavy Metals in the Environment: 147–215. NY: Marcel Dekker. Bhattacharya, A., Routh, J., Saraswathy, A., Jacks, G. & Bhattacharya, P. 2003. Influence of microbes on mobilization and speciation of arsenic from mine tailings in Adak, Västerbotten district, northern Sweden. 7th International Conference on Biogeochemistry of Trace Elements, Uppsala (Sweden), Vol. 2, pp. 58–59. Bhattacharya, A. 2004. Mobilisation of arsenic and other trace elements from abandoned mine tailings in Adak, Västerbotten district, northern Sweden. Licentiat thesis, Stockholm University. Cervantes, C., Ji, G., Ramirez, J.L. & Silver, S. 1994. Resistance to arsenic compounds in microorganisms. FEMS Microbiol. Rev. 15: 355–367. Collins, M.D., Routh, J., Saraswathy, A., Lawson, P.A., Schumann, P., Welinder-Olsson, C. & Falsen, C. 2004. Arsenicicoccus bolidensis gen. nov., sp. nov., a novel actinomycete isolated from contaminated lake sediment. Int. J. Syst. Evolut. Microbiol. 54: 605–608. Cullen, W.R. & Reimer, K.J. 1989. Arsenic speciation in the environment. Chem. Rev. 89: 713–764.

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Cummings, D.E., Caccavo, F.J., Fendorf, S. & Rosenzweig, R.F. 1999. Arsenic immobilization by dissimilatory Fe(III)-reducing bacterium Shewanella alga BrY. Environ. Sci. Tech. 33: 723–729. Diorio, C., Cai, J., Marmor, J., Shinder, R. & DuBow, M. 1995. An Escherichia coli chromosomal ars operon homolog is functional in arsenic detoxification and gram-negative bacteria. J. Bacteriol. 177: 2050–2056. Goriar, R., Chakraborty, K. & Pyne, R. 1984. Chronic arsenic poisoning from tubewell water. J. Indian Med. Assoc. 82: 34–35. Harvey, C.F., Swartz, C., Badruzzaman, A.B.M., Keon-Blute, N., Yu, W., Ali, M.A., Jay, J., Beckie, R., Niedan, V., Brabander, D., Oates, P., Ashfaque, K., Islam, S., Hemond, H. & Ahmed, M.F. 2003. Response to comments on arsenic mobility and groundwater extraction in Bangladesh. Science 300: 584. Islam, F.S., Gault, A.G., Boothman, C., Polya, D.A., Charnock, J.M., Chatterjee, D. & Lloyd, J.R. 2004. 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FEMS Letters 529: 86–92. Routh, J., Saraswathy, A., Nag, S.K., Sinha Ray, S.P. & Jacks, G. 2004. Arsenic reduction by indigenous bacteria in shallow aquifers from Ambikanagar, West Bengal (India). In Advances in Arsenic Research: American Chem. Soc. Special Issue (in press). Routh, J., Sinha Ray, S.P., Nag, S.K., Jacks, G., Bhattacharya, A., Datta, S. & Bhattacharya, P. 2003. Safe drinking water – The issue of shallow versus deep wells in two arsenic affected areas in West Bengal, India. 7th International Conference on Biogeochemistry of Trace Elements, Uppsala (Sweden), Vol. 2, pp. 462–463. Roy Chowdhury, T., Basu, G.K., Mandal, B.K., Biswas, R.K., Chowdhury, U.K., Chanda, C.R., Lodh, D., Roy, S.L., Saha, K.C., Roy, S., Kabir, S., Quamruzzaman, Q. & Chakraborti, D. 1999. Arsenic poisoning in the Ganges delta. Nature 401: 545–546. Saltikov, C.W. & Newman, D.K. 2003. Genetic identification of a respiratory arsenate reductase. PNAS 100: 10983–10988. Saha, K.C. 1984. 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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Geochemistry and geomicrobiology of arsenic in Holocene alluvial aquifers, USA J.A. Saunders, M.K. Lee & S. Mohammad Department of Geology and Geography, Auburn University, Auburn, USA

ABSTRACT: Groundwaters in Holocene alluvial aquifers in the USA contain elevated dissolved arsenic (As) have similar geochemical characteristics and microbiology. These include: (1) near-neutral pH and moderately reducing redox state; (2) significant dissolved iron and manganese; (3) presence of both iron- and sulfate-reducing bacteria; and (4) evidence that precipitation of authigenic carbonates and sulfides limits metal (loid) solubility. Our research indicates that anaerobic bacteria directly mediate dissolution of iron and manganese minerals in the alluvial aquifers leading to release of As and other trace elements, and also indirectly cause the precipitation of authigenic minerals such as siderite and rhodochrosite. Genetic sequencing indicated that the genus Geobacter are responsible for As release from Fe-oxyhydroxide. We propose a global-scale Geo-Bio-Hydro (GBH) arsenic cycle that directly invokes Pleistocene glaciers as important in physically weathering rocks setting the stage for Holocene chemical weathering that initiated a major release of As to the hydrosphere.

1

INTRODUCTION

There is now a general consensus that bacterial dissolution of arsenic-bearing iron oxyhydroxides (also known as hydrous ferric oxides, HFO) is the principal mechanism leading to As release in groundwaters hosted by Holocene river flood-plain deposits in such places as Bangladesh, India, Hungary, etc. (e.g. Chatterjee et al. 1995, Bhattacharya et al. 1997, Nickson et al. 1998, Acharrya et al, 2000, Nickson et al. 2000, Welch et al. 2000, McArthur et al. 2001, Smedley & Kinniburgh 2002, Dowling et al. 2002, Anawar et al. 2003, van Geen et al. 2003, Ahmed et al. 2004, Akai et al. 2004, Zheng et al. 2004). Further it is clear that this process occurs under moderately reducing redox conditions caused by the presence of aqueous or solid organic matter in the sediments (Nickson et al. 2000, McArthur et al. 2001). We (Saunders et al. 2003, Mohammad 2003) have proposed that the geochemical and microbiologic conditions leading to As-contamination of Holocene alluvial aquifers is universal in flood plains where river or stream sediments are deposited with organic matter. In this paper, we review our research findings from two field areas in the USA which appear to be in every way analogous to the problematic areas of the world adversely impacting human health due to As contamination. Further we show that one common type of iron reducing bacteria (FeRB) causes the As problem at of our USA field sites, where we have identified the “guilty” bacteria (present only in groundwaters with more than 50 ␮g/L As) using DNA sequencing techniques. Moreover, we propose this bacterium is likely responsible for the As problem elsewhere in the world, although clearly confirmation outside of USA is necessary and is planned. 2

GEOLOGY, GEOCHEMISTRY, AND GEOMICROBIOLOGY OF USA FIELD AREAS

Korte (1991) was first to document natural As contamination at an alluvial floodplain in the USA, and we chose his “discovery” area at the U.S. Department of Energy’s (DOE) Kansas City Plant (KCP) 155

to conduct our research. The KCP study area is located in floodplain deposits of the Blue River, a tributary of the Missouri River (Mohammad, 2003). The aquifer at the KCP site consists mainly of stream-valley alluvial deposits composed of Holocene alluvium which unconformably overlie Pennsylvanian strata of limestone and shale. Alluvial deposits typically consist of sand and gravel deposits of locally intermittent lenses of sand, silt, and clay. Groundwater at the KCP site is unconfined but locally can exist under confined conditions where clay layers are present (Mohammad 2003, Saunders et al. 1997). Water levels in wells at the KCP typically are ⬍7 m. Our second study area is located along Uphapee and Choctafaula creeks in an extensive Holocene floodplain deposits in central Alabama. These alluvial deposits have C-14 ages ⬍7000 years old and consist primarily of silty sand containing erratic but common macro wood fragments up 0.5 m in size. The alluvium unconformably overlies the Cretaceous Tuscaloosa Group of the Alabama Coastal Plain in the study area. Groundwater is unconfined with depths to the water table typically ⬍2 m. Groundwater at the Kansas City Plant is highly variable in redox state, iron (and manganese) content, and has near-neutral pH conditions. Arsenic was significantly elevated (e.g., ⬎50 ␮g/L in three of the eight monitoring wells, which also happened to have the lowest Eh values (Mohammad 2003). Dissolved Fe ranged from 0.2 to 17.1 mg/L, Mn from 0.5 to 7.9 mg/L, and alkalinity ranging from 175 to 350 mg/L (as CaCO3). Groundwater chemistry for the Alabama field site is generally similar to that of the KC site. It contains elevated Fe and Mn (up to 1 and 3 mg/L respectively, consistent with its moderately reducing nature), pH 6.6 to 6.8, 1–10 ␮g/L each of As, Co, Ni, Zn, REEs, and 50–175 ␮g/L of Ba. Further, sulfate reducing bacteria (SRB) in this alluvial aquifer use detrital wood fragments as the ultimate electron donor which leads to precipitation of pyrite containing up to 0.62 wt.% As and commonly, pyritization of the wood (Saunders et al. 1997). Several researchers have proposed that Fe- and possibly Mn-reducing bacteria may have been important in producing As-contaminated groundwater in Holocene alluvial aquifers. However, this has only been inferred from the high Fe (and Mn) contents of groundwaters with elevated As, its moderately reducing state, and a general correlation between dissolved Fe, As, and alkalinity (e.g., Nickson et al. 2000, McArthur et al. 2001, Ahmed et al. 2004). To evaluate this hypothesis, we conducted reconnaissance microbiologic investigations at our two field sites. At the Kansas City site, we collected two different types of samples for microbiologic investigations: groundwater for culturing of viable bacteria; and (2) bacteria filtered from several liters of groundwater in the field for molecular microbiology studies. Only wells at the Kansas City site that had elevated iron, Mn, and As could anaerobic bacteria be cultured, and both FeRB and SRB were found in these water samples. These groundwaters also had the lowest field Eh values. The microbiologic data suggest that FeRB are present and available to catalyze Fe-reduction in the As-enriched wells and thus this is the first direct indication from the field of their involvement in producing arsenic contamination in groundwater. Similar to the culturing experiments from the KCP site, cloning and sequencing of the 16S rDNA genes extracted from bacteria filtered from groundwater also indicated the presence of FeRB and SRB from the As-rich samples. Clones with sequences similar to known FeRB (e.g., Geobacter sp.) were also abundant in the As-rich water samples. Thus the molecular biological data corroborates the bacterial culturing results. In contrast to the Kansas City site, we used a truck-mounted auger to advance a borehole into the alluvial aquifer at the Alabama field site and collected solid samples below the water table for microbiologic and chemical analyses. Both FeRB and SRB were cultured from the aquifer, and enumeration using the most probable number (MPN) method indicated that FeRB were most abundant.

3

GEOCHEMICAL MODELING: IMPLICATIONS FOR ARSENIC MOBILITY

Geochemical modeling with PHREEQC (Parkhurst 1999) indicates that groundwater samples have positive saturation indices (SI) for siderite in two of the three As-elevated wells and supersaturated SI values for rhodochrosite in all three elevated-As wells at the KCP site. Similarly, all eight groundwater samples appear to be supersaturated with respect to goethite and understaurated 156

Figure 1. Redox-pH diagram for arsenic drawn at 25°C and fixed As and H2S activities of 10⫺ 2. Also shown is the groundwater geochemistry data from Bangladesh (oval) and the Kansas City Plant (cross ⫽ As ⬍ 10 ␮g/L; circle ⫽ 10–15 ␮g/L; diamond ⫽ ⬎50 ␮g/L).

with respect to ferrihydrite, suggesting that an iron oxyhydroxide phase with a solubility intermediate between those two might have been the source of iron in groundwaters. Additional geochemical modeling was conducted with the Geochemist’s Workbench (Bethke 1996) to compare groundwaters from the KCP study area to groundwaters from As-contaminated areas of Bangladesh, using published data from the British Geological Survey’s (BGS) “Special Study” where Eh was also measured in the field (British Geological Survey 2004). The BGS data plot in the same general area in Eh-pH space (Fig. 1) with a shift perhaps to slightly higher Eh values when the much larger data set is plotted. Similarly, Eh-pH plots were made for average geochemical conditions (with respect to Fe, Mn, carbonate, and S species) for both the KCP site groundwaters and those from Bangladesh (Fig. 1). KCP groundwaters plot near the Fe2⫹-siderite and Mn2⫹-rhodochrosite boundary suggesting they are approaching local equilibrium with both mineral phases (consistent with siderite SI values discussed previously). The most As-rich groundwater samples from the KC study area plot close to the stability field of pyrite (Fig. 1). As before, KCP and Bangladesh groundwaters plot in the same approximate areas in Eh-pH space. Similarly, groundwaters from Bangladesh appear to be close to equilibrium (or are supersaturated with) with respect to siderite and rhodochrosite. The Eh-pH values for the KCP waters indicate that they are o o in the stability fields of both arsenate (As-V, HAsO⫺ 4 ) and arsenite (As-III, As(OH)3 or H3AsO3), although the most As-enriched samples are in the arsenite stability field and even into the narrow stability fields of solid mineral phases realgar (AsS) and orpiment (As2S3). Results of our study support the general concept that As occurs in moderately reducing and typically Fe- and Mn-rich groundwaters. Anaerobic heterotrophic bacteria mediate the reductive dissolution of Fe and Mn minerals (and arsenic release) by the following chemical reaction: (1) where *As is sorbed As on iron oxyhydroxide and CH2O is generic organic carbon. Figure 1 shows that Mn-reduction occurs under more oxidizing conditions, and Mn-reducing bacteria will outcompete Fe-reducers as long as reactive solid Mn phases are available (Chapelle & Lovley, 1992). The presence and abundance of FeRB such as Geobacter in our two study areas provide the first 157

field evidence supporting the hypothesized bacterial reduction of iron oxyhydroxide and As release to groundwater. Moderately reducing, As-enriched groundwaters from the Kansas City site are generally supersaturated with respect to rhodochrosite and siderite. Saunders & Swann (1992) proposed that metabolism of Fe- and Mn-reducing bacteria could facilitate precipitation of both carbonate phases in aquifers because the products of reaction 1 become the reactants for Fe- and Mn-carbonate precipitation: (2) Mukherjee et al. (2001) and Pal et al. (2002) have observed authigenic siderite and rhodochrosite in alluvial aquifers in India, and documented textures very similar to those from Saunders & Swann (1992). Further, Saunders & Swann (1992) observed that authigenic carbonate phases often had inclusions of sulfide minerals such as pyrite and proposed that biogenic sulfate reduction probably occurred at the site of carbonate deposition. Because biogenic sulfate reduction raises pH and produces alkalinity as well as H2S (reaction 3a), this could explain both carbonate precipitation and iron sulfide formation (e.g., reaction 3b): (3a) (3b) Thus we propose that if Fe- and Mn-reducing bacteria are producing the moderately reducing, metal- and As-rich waters observed in this study and elsewhere in river floodplain deposits, then siderite and rhodochrosite are important phases controlling the solubility of iron and manganese in these waters. Precipitation of siderite will preferentially remove Fe relative to As in groundwater (Dowling et al. 2002), which can explain the often-observed poor statistical correlation for Fe and As in many similar groundwaters. By the same token, precipitation of iron sulfides could preferentially remove As relative to iron (by coprecipitation) and also affect the remaining As/Fe ratios in groundwater. Several researches have observed As-bearing pyrite as an authigenic phase in alluvial aquifers, which has led some researchers to propose that it was the source of As in alluvial aquifer groundwaters. This is highly unlikely due to the redox state of the water and the general lack of dissolved sulfate. We propose that SRB have removed limited sulfate in solution in these terrestrial waters and removed a small amount of Fe and As in the process (e.g., reactions 3a and 3b along with As coprecipitation in Fe-sulfide). Thus As-bearing pyrite is a sink for As rather than a source. 3

GEO-BIO-HYDRO (GBH) ARSENIC-CYCLING MODEL

We propose that a series of linked geochemical, biologic, and hydrologic processes operating over geologic time scales can lead to natural As contamination in Holocene fluvial and fluvial-glacial deposits around the world. We suspect that the conditions for this type of As contamination are very common, but the situation becomes particularly problematic in developing nations where existing surface water supplies are contaminated by parasites, fecal bacteria, etc. Further, our proposed GBH-As process requires neither As-rich (above crustal abundances ⬃2 ppm) source rocks nor rare or specialized microorganisms. The one “out-of-the-ordinary” aspect to the GBH process is the recognition that recent Earth history of continental glaciation and retreat exacerbates the natural As contamination process. “Normal” chemical weathering of minerals in folded mountain belts or other crustal rocks will release As to the hydrosphere and may even locally contaminate groundwater supplies. However, we propose that continental and/or alpine glaciation during the last ⬃5 million years lead to the present-day widespread occurrence of As in Holocene groundwaters in river floodplains that are most problematic around the world. Although clearly not all Ascontaminated groundwaters have been impacted by chemical weathering of glacial deposits, the 158

proximity of extensive As-contaminated groundwaters in Holocene river floodplains (e.g., in Indian Sub-continent, China, North America, Hungary) adjacent to significant glacial deposits “up-stream” argues for a genetic link in space and time. Low hydraulic gradients in the Holocene river floodplains exacerbates the problem, as the As-contaminated waters are not rapidly flushed out of these aquifers. Thus from the perspective of geologic time, the problem is ephemeral but unfortunately coincides with recent human history. The GBH process is initiated by continental glaciation and its associated physical weathering of continental rocks in its path. Glaciation sets the stage for accelerated chemical weathering of the reactive, high-surface-area minerals ground down by glaciers, which release As to the hydrosphere. Ferromagnesian silicate minerals typically contain minor amounts of arsenic and other trace elements. For example, the common sheet silicate biotite has a crystal structure leading to very rapid weathering (Saunders et al. 2000) and As release. Thus we propose that, Fe bearing silicates along with As-bearing sulfide and clay minerals are likely the ultimate source of As to the GBH-As process. Weathering of Fe bearing silicates and sulfides leads to the formation of iron oxyhydroxide, which are then physically transported by running water initially to streams and then through meandering stream channels. Some of the HFO would be colloidal in size and transported in suspension, larger grains, including HFO-coated sand grains, reside in stream bed load. Both suspended and bed-load HFO’s have a tremendous affinity to adsorb dissolved As. However, suspended HFO tend to move at the same average velocity as the transporting stream flow, and thus have a short residence time estimated to be ⬃2 weeks for major river systems discharging into oceans (Freeze & Cherry, 1979). On the other hand, HFO in river bed load has a much longer residence time and may become the loci of iron precipitation by iron-oxidizing bacteria in streams (Saunders et al. 1997); such biogeochemical reactions effectively create new and reactive HFO surfaces. The residence times of river bed-load sediment are difficult to quantify but small scale-studies suggest they may be in the order of 0.1 to 1 year per km length of the river (Meade 1982, Kelsey et al. 1987, Olley et al. 1997). Thus a major river system may have a water residence time of 2 weeks and a bed load residence time of 103 to 104 years. So using a residence time of 5 ⫻ 103 years for a major river system, stream bed load would have ⬃1.3 ⫻ 105 river volumes flowing over and through it during its transport in the river, and As is slowly adsorbed onto the HFO surfaces. We propose that the long-term adsorption is the important As-concentration mechanism that ultimately leads to natural As contamination of groundwater in the lower reaches of the river systems where extensive floodplains develop. In the floodplains, reactive organic matters co-deposited with As-bearing HFO’s causes anaerobic conditions and FeRB metabolism. This portion of the GBH cycle is essentially the “HFO-As” hypothesis that we and others have proposed previously. We propose that the HFO-As hypothesis is just a part of a much bigger continent-scale process. Fluvial and fluvial-glacial deposits typically have low dissolved sulfate as opposed to river deltas where seawater sulfate can mix with fresh groundwater, which has caused the formation “high sulfur” (and locally As?) Carboniferous coal deposits in the eastern USA and elsewhere. Thus SRB cannot effectively remove a significant amount of As released by FeRB bacteria in the anaerobic flood-plain groundwater. However SRB do cause some local pyrite precipitation and As removal (Saunders et al. 1997, Nickson et al. 2000, McArthur et al. 2001); this has led to confusion about pyrite being a source of As in river flood-plain sediments. Pyrite is really an As sink that also causes the observed inverse correlation between dissolved As and sulfate seen in Bangladesh and our Kansas City study area. A major implication of the GBH-As cycle is that it leads to a prediction of potential problem areas for natural As contamination as world population increases and new water sources are required. Currently, because groundwater in Holocene aquifers have elevated iron contents, it has led to the under usage of these groundwaters in the developed nations, and many of these same groundwaters may also happen to have elevated As.

4

POSSIBILITY OF BIOREMDIATING ARSENIC USING SRB

We have successfully stimulated naturally occurring SRB to remediate metal-contaminated groundwater at a number of sites (Saunders et al. 2001, Lee & Saunders 2003, Saunders et al. 2004). SRB 159

metabolism causes certain toxic metals to precipitate out of groundwater as relatively insoluble sulfide minerals (e.g., Zn, Pb, Cd, Ag, Cu, Hg, etc.), some redox-sensitive metals to precipitate as oxy-hydroxide phases (e.g., Cr, U), and raises groundwater pH which can also lead to increased sorption of some elements. However, raising pH can cause As to desorb. This, coupled with the possibility of As-thio complexing (Wilkin et al. 2003), makes the geochemical behavior of As under sulfate-reducing conditions not so straightforward even though it forms its own sulfide minerals and coprecipitates in Fe-S phases. However, our research suggests that these problems can be overcome if a strategy of optimizing the geochemical conditions for removal of As from solution is employed. It appears that high dissolved iron content favors arsenic coprecipitation in Fe-S sulfide phases under sulfate-reducing conditions, and limits the As-thio complexes (Wilkin et al. 2003). Thus, we propose that bioremediating As-contaminated groundwater is possible by adding a solution of hydrous ferrous sulfate and a carbon electron donor (e.g., sucrose, molasses, methanol, ethanol, etc.) through injection wells. The lack of ferrous iron can limit SRB metabolic efficiency because Fe(II) is present in many of the enzymes and compounds used in electron transfers by SRB. In the past, researchers have assumed that H2S was toxic to SRB, but actually the “toxicity” effect of H2S is the removal of Fe(II) by Fe-S precipitation. Thus, injection of dissolved Fe(II) into As-contaminated groundwater: (1) insures that iron will be available for SRB metabolism; (2) limits the buildup of potentially toxic levels of H2S; (3) keeps the Fe(II)/H2S ratio high enough to keep As-thio complexes from occurring to any significant extent, and (4) provides both the Fe and S needed for the As-“encapsulating” FeS phases. Thus this approach can be effective in treating Ascontaminated groundwater as the conditions for As-coprecipitation in FeS are optimized.

4

CONCLUSIONS

Regulatory agencies around the world typically have a “secondary” drinking water standard (not health-based) for iron of 0.5 mg/L, and thus Fe-rich groundwater in the United States and other “developed” nations has largely been used for irrigation of crops and not for human consumption. We believe the high iron content has led to the under-appreciation of the fact that elevated As is common in alluvial aquifers in the developed nations of the world as well as countries such as a Bangladesh, India, Pakistan, etc. The major difference is that rivers in the developed nations are kept clean enough for drinking water purposes because of the investment in expensive infrastructure, and rivers in developing nations are commonly polluted and not fit for drinking. This has driven the use of the iron- and locally As-rich groundwaters of Holocene river floodplains. Previous studies have noted that groundwaters with elevated arsenic in alluvial floodplain deposits also have elevated iron (typically 1–50 mg/L). Other specific conclusions: (1) We have identified Geobacter using DNA analysis (and FeRB by culturing) in groundwaters with elevated Fe, Mn, and As in our USA field areas and we propose that this type of anaerobic FeRB bacteria cause the As problem in Holocene alluvial aquifers around the world, but verification is needed. (2) Authigenic carbonate and sulfide minerals appear to precipitate as a consequence of anaerobic bacteria metabolism. These minerals will affect observed present-day groundwater As/Fe ratios and metal solubility. (3) Our GBH-As cycle proposes that Pleistocene glaciers played a major role in the weatheringinduced release of As to the hydrosphere in the Holocene, and that the As-concentrating mechanism is sorption by HFO in stream sediments which have much longer residence time than river waters. The subsequent action of dissimilatory iron and sulfate anaerobic bacteria control the fate of As and other redox-sensitive elements in Holocene alluvial aquifers. (4) Bioremediation may be possible using SRB if the Fe/H2S ratio is kept elevated in As-bearing groundwater to prevent the formation of soluble thio-arsenite species. However, this is not likely a panacea for remediating natural As contamination in the Bengal Basin and similar areas, but may prove useful for anthropogenic arsenic pollution. 160

REFERENCES Acharyya, S.K., Lahiri, S., Raymahashay, B.C. & Bhowmik, A. 2000. Arsenic toxicity of groundwater in parts of the Bengal Basin in India and Bangladesh; the role of Quaternary stratigraphy and Holocene sea-level fluctuation. Environ. Geol. 39: 1127–1137. Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A.H., Imam, M.B., Khan, A.A. & Sracek, O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh: An overview. Appl. Geochem. 19: 181–200. Akai, J., Izumi, K., Fukuhara, H., Masuda, H., Nakano, S., Yoshimura, T., Ohfuji, H., Anawar, H.M. & Akai, K. 2004. Mineralogic and geomicrobiologic investigations on groundwater arsenic enrichment in Bangladesh. Appl. Geochem. 19: 215–230. Anawar, H.M., Akai, J., Komaki, K., Terao, H., Yoshioka, T., Ishizuka, T., Safiullah, S. & Kato, K. 2003. Geochemical occurrence of arsenic in groundwater of Bangladesh; sources and mobilization processes. J. Geochem. Explor. 77: 109–131. Bhattacharya, P., Chatterjee, D. & Jacks, G. 1997. Occurrence of arsenic-contaminated groundwater in alluvial aquifers from Delta Plain, Eastern India: options for safe drinking water supply. Wat. Res. Dev. 13: 79–92. Bethke, C.M. 1996. Geochemical Reaction Modeling. New York, Oxford University Press. British Geological Survey 2004. The groundwater arsenic problems in Bangladesh (phase 2), website: www.bgs.ac.uk/arsenic/bangladesh/Data/SpecialStudyData/csv. Chapelle, F.H. & Lovely, D.R. 1992. Competitive exclusion of sulfate reduction by Fe (III)-reducing bacteria. Ground Water 30: 29–36. Chatterjee, A., Dipankar Das, Mandal, B.K., Chowdhury, T.R., Samanta, G. & Chakraborti, D. 1995. Arsenic in Ground Water in Six Districts of West Bengal, India. The Biggest Arsenic Calamity in the World. Part 1. Arsenic Species in drinking water and urine of the affected people. The Analyst 120: 643–650. Dowling, C.B., Poreda, R.J., Basu, A.R. & Peters, S.L. 2002. Geochemical study of arsenic release mechanisms in the Bengal Basin groundwaters. Wat. Resour. Res. 38: 1173–1190. Freeze, R.A. & Cherry, J.A. 1979. Groundwater. New Jersey, Prentice-Hall. Kelsey, H.M., Madej, M.A. & Lamberson, R. 1987. Stochastic model for the long term transport of stored sediment in a river channel. Wat. Resour. Res. 23: 1738–1750. Korte, N.E. 1991. Naturally occurring arsenic in groundwaters of the Midwestern United States. Environ. Geol. & Wat. Sci. 18. Lee, M.-K. & Saunders, J.A. 2003. Effects of pH on metals precipitation and sorption: Field Bioremediation and geochemical modeling approaches. Vadose Zone J. 2: 177–185. McArthur, J.M., Ravenscroft, P., Safiulla, S. & Thirlwall, M.F. 2001. Arsenic in groundwater; testing pollution mechanisms for sedimentary aquifers in Bangladesh. Wat. Resour. Res. 37: 109–118. Meade, R.H. 1982. Sources, sinks, and storage of river sediment in the Atlantic drainage of the United States. J. Geol. 90: 235–252. Mohammad, S. 2003. Universality of geochemical, microbiologic, and hydrogeologic processes controlling the development of natural arsenic contamination of Holocene flood-plain aquifers. M.S. Thesis. Auburn University. Mukherjee, P.K., Pal, T., Sengupta, S., & Shomw, S. 2001. Arsenic rich phases from aquifer sediments from southern West Bengal. J. Geol. Soc. India 58: 173–176. Nickson, R., McArthur, J., Burgess, W., Ahmed, K.M., Ravenscroft, P. & Rahman, M. 1998. Arsenic poisoning of Bangladesh groundwater. Nature 395: 338. Nickson, R.T., McArthur, J.M., Ravenscroft, P., Burgess, W.G. & Ahmed, K.M. 2000. Mechanism of arsenic release to groundwater, Bangladesh and West Bengal. Appl. Geochem. 15: 403–413. Olley, J.M., Roberts, R.G. & Murray, A.S. 1997. A novel method for determining residence times of river and lake sediments based on disequilibrium in the thorium decay series. Wat. Resour. Res. 33: 1319 Pal, T., Mukherjee, P.K., Sengupta, S., Bhattacharyya, A.K. & Shome, S. 2002. Arsenic pollution in groundwater of West Bengal, India; an insight into the problem by subsurface sediment analysis. Gondwana Res. 5: 501–512. Parkhurst, D.L. & Appelo, C.A.J. 1999. User’s guide to PHREEQC (Version 2)—A computer program for speciation, batch-reaction, one-dimensional transport, and inverse geochemical calculations. U.S. Geol. Surv. Water-Resour. Inv. Report 99(4259): 310. Saunders, J.A. & Swan, C.T. 1992. Nature and origin of authigenic rhodochrosite and siderite from the Paleozoic aquifer, northeast Mississippi, U.S.A. Appl. Geochem. 7: 375–387. Saunders, J.A., Pritchett, M.A. & Cook, R.B. 1997. Geochemistry of biogenic pyrite and ferromanganese stream coatings: A bacterial connection? Geomicrobiol. J. 14: 203–217.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Arsenic contamination in drinking water of tube wells in Bangladesh: statistical analysis and associated factors M. Anwar Hossain Department of Farm Structure, Faculty of Agricultural Engineering & Technology, Bangladesh Agricultural University, Mymensingh, Bangladesh

M. Amirul Islam Department of Agricultural Statistics, Faculty of Agricultural Economics & Rural Sociology, Bangladesh Agricultural University, Mymensingh, Bangladesh

M.O. Gani & M.A. Karim Graduate Student, Faculty of Agricultural Engineering & Technology, Bangladesh Agricultural University, Mymensingh, Bangladesh

ABSTRACT: Recently arsenic contamination in tubewell water is of a growing concern among the policy makers and researchers. This study has been devoted to the investigation of the existing arsenic contamination level in tubewell water and as well as the association among different characteristics of tubewells of two-selected regions of Bangladesh. The data were collected from a highly arsenic contaminated area at Chuddyagram Upazila in Comilla district and from a low arsenic contaminated area at Modhupur Upazila in Tangail district. More than two third of the tubewells of that region are private and 53.3% of the tubewells have been installed in the last ten years. 27.1% of the tubewells were found to be used by 11–20 persons regularly. Around 35.9% of the tubewells were completely unaffected by arsenic. About 68.8% of the tubewells have been installed above 30 m depth level. Among the tubewells under study 66.3% were found to be marked as safe. Among the users of the tubewells, 25.7% were found to be under risk of arsenic pollution by tubewell water. Most of the government tubewells (83.8%) are safe, whereas 58% of the private tubewells are arsenic free. Multiple linear regression analysis suggested that both the depth and age of installation of tubewells have significant negative impact on the level of arsenic. Number of users was found to be significantly influenced by depth and age of installation in positive direction and by arsenic contamination level in negative direction. Linear logistic regression has been used to identify the factors that significantly affect the status of a tubewells to be identified as arsenic contaminated and only depth level of tubewells was found to have significant negative effect on the dependent variable. Finally, on the basis of the findings a set of recommendations have been suggested in this article, which will help the policy makers to take future initiatives to overcome the problem.

1

INTRODUCTION

Since the last decade Bangladesh has been facing an increasing rate of arsenic contamination in ground water, which has resulted in a serious threat to public health through potable water from tubewells. About 57 million people (out of the total population of 140 million) are in the grip of arsenic contamination from shallow tubewells originally installed to overcome the shortage of drinking water. The nature and extent of arsenic contamination in rural villages, where about 76% people are living depending on the supply of tubewell water from shallow aquifers, which have been known to be arsenic contaminated, seem severe. Arsenic in shallow tubewell water has been 163

detected in almost all parts of Bangladesh. In the acute arsenic problem areas, more than 90 percent of the shallow tubewells have been found to yield contaminated water exceeding 50 ␮g/L of arsenic. The WHO established 10 ␮g/L as a Provisional Guideline Value for arsenic in 1993 (WHO 2000). The parliament recently informed that arsenic has been detected in ground water in 59 out of 64 districts of the country. The districts where arsenic was not found were northern Jamalpur, three hill districts of Rangamati, Khagrachhari, Bandarban and the southern most Cox’s Bazar (Bashar 1999). A significant percentage of the existing 10 million tubewells in Bangladesh is contaminated with arsenic. WHO has estimated that the death of 0.2 to 0.27 million people has occurred from cancers induced due to the elevated concentrations of arsenic in groundwater he consumption of high arsenic (The Bangladesh Observer, June 25, 2001). Arsenic contamination in the ground water appears to be a major problem in Bangladesh. About 97% of the population of Bangladesh use tubewell water for drinking and cooking purposes (DCH 1998) but the arsenic contaminated ground water is now major environmental hazard and natural problem of the country. Nearly 20–22 million people in Bangladesh are currently consuming ground water contaminated with arsenic. Another 70 to 75 million people are at risk (Mortoza 2001). The Department of Public Health Engineering (DPHE) initiated monitoring and water testing activities in the adjacent border districts in 1993. The WHO recommended level for arsenic in drinking water at 50 ␮g/L is now considered by scientists to be far too high. The revised level at 10 ␮g/L may also be too high for a warm country like Bangladesh and if true, a far higher number of the people than the 20 to 22 million quoted will be drinking water containing high levels of arsenic. Excessive level of arsenic in body can lead to some serious health hazards as melanosis, leucomelanosis, keratosis, hyperkeratosis, dorsum, non-petting cedema, gangrene, skin cancer, hyper pigmentation, depigmentation, and a number of internal cancers. Cardiovascular and neurological diseases have also been found to be linked to the arsenic contamination. Elevated concentration of arsenic in water when associated with malnutrition and hepatitis B, which are most common in Bangladesh, accelerate the effects of arsenic poisoning. Over the last 25 years the number of people drinking arsenic contaminated water has increased due to well drilling and population growth in Bangladesh. The number of affected persons may, therefore, increase further. The most important remedial measure is prevention of further exposure by providing safe drinking water. Efforts have been taken in this article to examine in what extent arsenic level is influenced by some selected variables and thus suggest some policy measures to overcome the threatening situation. 2

METHODOLOGY

This study is based on the data gathered from Thana Public Health Engineering Department. The primary data was collected nation wide by using a questionnaire through the national survey carried out under the Bangladesh Arsenic Mitigation and Water Supply Project (BAMWASP) and the DPHE. We considered the information about tubewells from two selected areas: i) high arsenic affected Chuddyagram Upazila in Comilla district, and ii) low arsenic affected Modhupur Upazila in Tangail district. The data consists of information about 704 tubewells in these selected areas. Multivariate analyses e.g., multiple linear regression and binary logistic regression have been used to explore the associations of some characteristics of tubewells with arsenic contamination level in the water from these tubewells. 2.1

Multiple linear regression

To understand the nature of relationship among different tubewell characteristics multiple linear regression analysis was considered to fit the regression models of the form: (1) 164

Where, Y ⫽ dependent variable, a ⫽ intersect and bi ⫽ partial regression co-efficient associated with the independent variable Xi and N is the number of independent variables in the model. In the multiple linear regression analysis we have considered ‘arsenic contamination level’ and ‘number of users’ as dependent variables with ‘age of installation’ and ‘depth of tubewell’ as independent variables. 2.2

Linear logistic regression

When the dependent variable is qualitative with two levels or categorised into two groups this type of regression can be used. Let Yi denote the dependent variable for the ith observation and Yi ⫽ 1 if the ith individual is a success and Yi ⫽ 0, if the ith individual is a failure. Suppose for each of the n individuals, k independent variables Xi1, Xi2 …., Xik are measured. In the linear logistic model as suggested by Cox (1970), the dependence of the probability of success of independent variables is assumed to be

(2)

(3)

where, Xi0 ⫽ 1 and bj’s are the unknown regression coefficients and j ⫽ 0, 1, 2 … .k. From the above two equations, we get,

(4)

Equation (4) expresses the linear logistic regression model in which the parameters are estimated by maximum likelihood method.

3 3.1

RESULTS AND DISCUSSION Background characteristics

Table 1 reveals the distribution of tubewells by different characteristics. About 68 percent of the tubewells are private and the rest are installed by the governmental agencies. More than half (53.3%) of the tubewells have been installed during the last ten years. This is followed by the installation age 11–20 years (27.7%). A few among the tubewells (0.1%) are still in use with more than 40 years of installation age. Around sixty-nine (68.8%) percent of the tubewells have been installed at 30 m depth level or less, which is followed by the depth level 31–45 m (29.5%). About 52% of the tubewells have been used by 1–10 persons. This percentage decreases with the increase in number of users. Ten percent of the tubewells were used by more than 30 people. Around 35.90% of the tubewells were completely unaffected by arsenic, whereas arsenic contamination within the safe range (10–49 ␮g/L) have been found for 30.4% of the tubewells. About 33.6% of the tubewells were found seriously contaminated with arsenic. All the tubewells under study have 165

Table 1.

Distribution of tubewells by different characteristics.

Characteristics

Percentage

Type of ownership Government Private Age of installation (years) 0–10 11–20 21–30 31–40 41–50 Depth of tubewells (m) ⭐30 31–45 46–60 Number of users 1–10 11–20 21–30 31–40 ⬎40 Arsenic contamination level (␮g/L) ⬍10 10–49 50–100 101–250 Arsenic identified Green wells with no arsenic Red wells with arsenic

Table 2.

32.5 67.5 53.3 27.7 15.8 3.1 0.1 68.8 29.5 1.7 51.5 27.1 11.2 4.4 5.8 35.9 30.4 33.2 0.4 66.3 33.7

Distribution of tubewell users.

Arsenic status

Percentage of users

Average users per tubewells

Green wells with no arsenic Red wells with arsenic

74.3 25.7

17.6 (sd. ⫽ 14.8) 11.9 (sd. ⫽ 24.1)

been checked by field workers of the BAMWASP for possible presence of arsenic. The arsenicsafe wells were marked with green color while those seriously contaminated by arsenic were painted red with suggestion to avoid these well for drinking purposes. Among the tubewells under study 66.3% were found to be marked as safe, the remaining wells are restricted for use (Table 1). Among the users of the tubewells 25.7% were found to be under risk of arsenic pollution by tubewell water (Table 2). On average, about 12 persons of the study area are using arsenic contaminated tubewells water, whereas 18 persons on average are the users of safe tubewells. 3.2

Bivariate analysis

Relationships among different variables in terms of the Pearson’s correlation coefficient are presented in Table 3. All of the variables e.g. depth of tubewell, number of users, arsenic contamination level and age of installation are significantly correlated with each other. The relation of depth level of tubewell with number of users and age of installation are positive and that with arsenic contamination level is negative. This suggest that if the possible number of user is large, the depth of the tubewell should be planned as deep accordingly. The tubewells, those are of comparatively 166

Table 3.

Correlation coefficient among four selected variables.

Name of variable

Depth of tubewells

Number of users

Arsenic level

Age of installation

Depth of tubewells Number of users Arsenic level Age of installation

1.000 0.295*** ⫺0.254*** 0.258***

0.295*** 1.000 ⫺0.158*** 0.310***

⫺0.254*** ⫺0.158*** 1.000 ⫺0.139***

0.258*** 0.310*** ⫺0.139*** 1.000

Note: Significance level *** P ⬍ 0.01.

Figure 1.

Distribution of arsenic levels with the depth of the tubewells.

higher depth have possibility to supply ground water for longer period. The negative relation of depth of the wells with the arsenic contamination level reveals that as depth is increased the arsenic contamination level is supposed to decrease (see Fig. 1). This is an important phenomena that was also observed by National Hydrochemical Survey Report, 2000. Number of users has a negative relation with arsenic contamination level indicating that people use to avoid arsenic contaminated tubewells if there is any alternative. The positive relation of number of users with the age of installation does not necessarily mean that with the increase of this parameter, the other parameter will increase at the same time. Rather it reveals the fact that if a tubewell is found to be safe and comfortable in terms of all the criteria, there is a least chance to avoid it. With a minor maintenance, the outflow of water from that tubewell is kept continuous, hence increasing its longevity (its age from installation) and attracts more people from an accessible surrounding. With a similar manner the negative relation of arsenic contamination level and age of installation can be interpreted; if a tubewell is found as contaminated with arsenic, people try to avoid it (some times to abandon completely) if there is any alternative, hence the lasting age (exploitation period) remains minimum. On the other hand, for a safe tubewell people use to maintain all safety measure to keep the tubewell working. 3.3

Differences in tubewell ownership

Possible relationship of ownership of a tubewell with some variables, e.g., arsenic found, depth level, number of users and age of installation have been judged by the Pearson’s chi-square test (Table 4). Most of the government tubewells (83.80%) are safe, whereas, 58.0 percent of the 167

Table 4.

Distribution of tubewell ownership by some selected variables. Type of ownership

Variables Arsenic identified No Yes Depth level (m) ⭐30 31–45 46–60 No. of users per tubewells 1–10 11–20 21–30 31–40 41 and more Age of installation (years) 0–10 11–20 21–30 31–40 41–50

Government (%)

Private (%)

Chi-square value

83.8 16.2

58.0 42.0

45.9***

26.3 71.1 2.6

89.1 9.7 1.2

286.7***

20.2 33.3 21.9 10.1 14.5

66.6 24.2 6.1 1.7 1.5

174.3***

26.3 34.6 31.1 7.5 0.40

66.2 24.0 8.0 1.1 —

124.7***

Note: Level of significance *** P ⬍ 0.01.

private tubewells are arsenic free. Government initiated tubewells were mostly installed at depth levels of 31–45 m (among 71.1%), on the other hand most of the private tubewells (89.1%). are installed at depths of 30 m or less. In private sector people do not care about design for tubewells; and rely on the method of tubewell installation, which is most economic at shallow depths. The Department of Public Health Engineering however follow a set up design for tubewell installation on the basis of geological characteristics of aquifer of a specific region. Thus, they use to install a tubewell in a greater depth to get sufficient good quality potable water. These findings are also supported by the National Hydrochemical Survey Report on Ground Water Pollution in Bangladesh (NHCSR 2000). Tubewells installed by governmental organisations are set to serve comparatively big user population. About 33.3% percent of the tubewells are used by 11–20 users every day. This is followed by the group ranging from 21–30 users (21.9%), 1–10 users (20.2%), 41 and more users (14.5%) and 31–40 users (10.1%). For most of the private tubewells, targeted number of users is 1–10 (66.6%). The second level is 11–20 (24.2%). About three fourth of the Government tubewells are more than ten years old. In the contrast 66.2% private tubewells are less than 10 years old. One fourth of the private tubewells are within the age range of 11–20 years (24.4%). All of these variables (or categorised variables) are found to be significantly associated with ownership of a tubewell. Table 4 indicates that in all the respects Government tubewells and private tubewells possess completely different characteristics, which may be the basis for the difference in arsenic contamination level in the tubewells. 3.4

Multivariate statistics

Multiple linear regression has been used to identify the relationship of arsenic contamination level of tubewell water (Y) with depth level of tubewell (X1) and age of installation of tubewell (X2) (Table 5). It reveals from the table that the resulting model is highly significant (overall and individual coefficient). Both the coefficients with negative sign indicate that with the increase in depth level or age of installation, the arsenic contamination level decreases. This again supports the 168

Table 5. Results of the Multiple Linear Regression of arsenic contamination level on depth level and age of installation of tubewells. Independent variables

Coefficient

t

F

R2

Depth level Age of installation Constant

⫺0.0002631*** ⫺0.0003014** 0.05347***

⫺6.211 ⫺2.089 14.415

26.59

0.071

Note: Level of significance *** P ⬍ 0.01 and ** P ⬍ 0.05; Significance level 0.0000.

Table 6. Results of the Multiple Linear Regression of number of users of tubewells with some selected variables. Independent variables

Coefficient

t

F

R2

Arsenic contamination level Depth level Age of installation Constant

⫺39.215* 0.136*** 0.529*** ⫺0.741

⫺1.922 5.786 6.761 0.325

41.173

0.150

Note: Level of significance *** P ⬍ 0.01 and * P ⬍ 0.10; Significance level 0.000.

result in the correlation matrix (Table 3) along with the interpretations stated earlier. Finally the mode becomes (5) The relationship of number of users of tubewells (Y) with arsenic contamination level (X1), depth level of tubewells (X2) and age of installation of tubewells (X3) was also investigated (Table 6). A multiple linear regression with a statistically significant model has been estimated, with all the independent variables having significant relation with the dependent variable. Number of users has positive relation with depth level and age of installation indicating that number of users will increase with the increase of these two variable values. Negative effect of arsenic contamination level is explained by the general attitude of people that a user will avoid the arsenic affected tubewell. The interpretations behind positive relation are already discussed on the basis of correlation table (Table 3). The model can be represented as: (6) Since the sample size is large, possible multi-colinearity among the independent variables can be ignored.

3.5

Logistic Regression Analysis

Finally, an effort has been made to judge the impact of different variables on the status of the tubewells as arsenic identified or not, using linear logistic regression (Cox 1970, Fox 1984), taking ‘Arsenic identified’ (0 ⫽ No. 1 ⫽ Yes) as dependent variable (Table 7). Depth level has significant negative effect on the status of a tubewell to be arsenic contaminated or not. The result suggests that as the depth level decreases the chance of the tubewell to be identified as arsenic affected increases. This supports earlier discussion about the relation between depth and arsenic contamination, presented in the previous section of this article. 169

Table 7. Results of the Linear Logistic Regression of arsenic contamination and different independent variables. Independent variable

Level

Coefficient

Depth level Continuous ⫺0.021*** Age of installation Continuous 0.013 Number of users Continuous 0.012 Constant – 1.282*** Model chi-square ⫽ 66.44; d.f.: 3; Significance level. 0.000 Note: Level of significance: *** P ⬍ 0.01

4

CONCLUSION

Arsenic contamination in drinking water has already appeared as a major threat to the public health in Bangladesh. As a quick response researchers and policy makers have already taken the matter as a priority issue. The problem is deep rooted and most of which is a consequence of unplanned man-made activities. Withdrawal of water from rivers in the Indian Territory, continuous deforestation and excessive use of ground water has already initiated an unaffordable climatic/geological change in Bangladesh. Among other dangerous outcomes of these activities, arsenic contamination is one of the most vital threats. There is an emerging risk of arsenic intake through agricultural products, besides through drinking water from the underground source. In this article we investigated only the contamination of tubewell water by arsenic. Arsenic intake through drinking water is more serious than any other exposure pathways and for a large number of people of Bangladesh tubewells are the only source of drinking water. There has been a lot of research focused on arsenic removal from water. So far none of the findings and suggestions is still affordable at a large scale due to the excessive cost involvement. Until there is an acceptable cost-effective, easily available alternative, we should focus on the modification and proper planning of the existing tubewell installation system. This paper identifies some significant relationships of arsenic contamination level with some tubewell characteristics. If there could be an adjustment of the tubewell parameters before installation of the tubewells, the chance of arsenic contamination may be reduced. A major planning difference has been identified between Government and private initiative regarding the depth level of tubewells. Considering the relation of arsenic contamination level with depth level, local governments may be authorized to monitor the depth level of installation. This authority should also update information on the overall arsenic contamination level, including a safe depth level to install a tubewell along with other geological data of that region. There should also be a dissemination of information program run for all time to inform and motivate the inhabitants to be careful in installing a tubewell rather than minimizing its cost. None of the scientific effort can eradicate arsenic forever. The only way to get rid of the problem permanently is to restore the natural and climatic condition, as it should be for a peaceful and safe environment. Before finalizing a most acceptable solution, effort to minimize the probable arsenic intake will be beneficial all over the country. Government should also think about giving subsidy for the installation of tubewells where necessary. This paper has been devoted to assess the existing tubewell installation scenarios and outcomes with the suggestions and hope that at least we are able to minimize the risk only by proper design and monitoring process. Community and management level awareness and initiatives are the primary steps toward solution of the problem.

REFERENCES Bashar, A. 1999. Arsenic contamination mitigation efforts. The Observer Magazine, 16th April 1999. Cox, D.R. 1970. Analysis of Binary Data, Chapman and Hall. London, UK.

170

D.C.H., 1998. Dhaka Community Hospital. The Arsenic Problem in Bangladesh: An Introduction to the country and its arsenic situation. Presented at Int. Conf. On Arsenic Pollution of Ground Water in Bangladesh. 8–12 Feb., 1998, Dhaka. Abstract. pp. 27–30. Fox, J. 1984. Linear Statistical Modes and Related Methods, John Wiley and Sons, New York, USA. Mortoza, S. 2001. Ground Water Contamination with Arsenic. The Bangladesh Observer, 2001. NHCSR 2000. A Executive Summary Main Report of National Hydro-Chemical Survey on Groundwater Studies of Arsenic Contamination in Bangladesh by British Geological Survey and Mott MacDonald (UK) for Govt. Bangladesh, Min. Local Govt Rural Devel. Cooperatives, Dept. Public Health Engrg., and for Dept. Intl. Devel. (DFID-UK)., Dhaka, Bangladesh. WHO 2000. World Health Organization Sustainable Development and Healthy Environments “Towards an Assessment of the Socioeconomic Impact of Arsenic Poisoning in Bangladesh” Geneva, Switzerland.

171

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

The impact of low dissolved oxygen in recharge water on arsenic pollution in groundwater of Bangladesh Md. Nazrul Islam Department of Civil Engineering, University of Toronto, Toronto, Canada

R.D. Von Bernuth Department of Biosystems Engineering, Michigan State University, East Lansing, Michigan, USA

ABSTRACT: This study is based on the concept that lack of dissolved oxygen (DO) at or below the water table and water extraction (Q) through shallow irrigation wells at a rate greater than the aquifer recharge rate are the main causes of arsenic release in the groundwater of Bangladesh. This study identified the hydrogeochemical processes related to shortage of DO that eventually produce high arsenic concentrations and their migration into the groundwater systems. The existing theories of arsenic release by oxidation and reduction in the context of dissolved oxygen shortage in recharging groundwater were studied. Both numerical and thermodynamic analyses were used to demonstrate how oxidation theory of arsenic release is inadequate to explain the release of arsenic into the groundwater of Bangladesh. This study quantified the amount of dissolved oxygen level in deeper layers of the aquifer and their relation to the variations in redox potential values and arsenic release processes. It also analyzed groundwater velocity and flow patterns to establish a link between dissolved oxygen shortage and arsenic release into the groundwater. On the basis of the findings, it was concluded that shortage of dissolved oxygen in recharging water is the most likely the root cause of arsenic occurrence in Bangladesh groundwater.

1

INTRODUCTION

The arsenic contamination problem in Bangladesh groundwater is most likely to be associated with shortage of dissolved oxygen in recharge water at or below the water table. The presence or absence of dissolved oxygen (DO) in natural surface or groundwater systems has been known as an index of Oxidation Reduction Potential (ORP) or redox potential. The ORP is a measure of tendency for donating or accepting electrons during chemical reactions. As long as any measurable amount of dissolved oxygen is present in the groundwater systems, the redox potential (pe) level is controlled by the dissolved oxygen concentration and the system will remain in oxic conditions. Under higher oxic conditions, ion activities and electron donating tendencies are less (Stumm & Morgan 1996, Khan et al. 2000). On the contrary, under anoxic conditions, the tendency for donating electrons is high. Arsenic concentrations in the groundwater in Bangladesh are a by-product of redox reactions where microbial derived oxidation of organic carbon plays an important role in donating electron, and the terminal electrons are accepted by the hydrated ferric oxides and/or hydroxides present in the groundwater systems. Upon electron acceptance, ferric oxides are reduced from solid particulate form to dissolved ferrous ions and the associated arsenic is released into the groundwater systems (Nickson et al. 2000). This is the most widely accepted reduction theory of arsenic release into the groundwater of Bangladesh. The origin of arsenic in the aquifer sediment is from natural geological processes. There are two theories of arsenic release. One is that over time the microbial degradation of organic carbon has caused arsenic to be released into water by reductive dissolution of iron oxides under anoxic or 173

mildly reduced conditions. This is known as reduction theory of arsenic release. However, some experts believe that exposure of arsenic rich pyrites to the atmospheric oxygen due to over extraction of groundwater is the main cause of arsenic release into the groundwater of Bangladesh (Bridge & Hussain 1999). This idea is known as the oxidation theory of arsenic release. Neither of these two theories is unequivocally accepted by all and a strong debate has been continuing on these two current theories of arsenic release. It is not the purpose of this research to engage in this debate. In both cases dissolved oxygen level plays an important role in arsenic contamination problem, and that is the focus of this study. The ORP or redox potential of the sediment water systems mainly controls the arsenic speciation in groundwater systems (Bhattacharya et al. 2002a b, Ahmed et al. 2004). The Speciation or different oxidation states of arsenic species in water is highly attributed to the physicochemical (electrostatic force, ionic strength, pH) and molecular interactions (thermodynamic stability, water solubility, hydrogen bonding ability etc.) (Gazsó 2001). The physicochemical and molecular interactions between arsenic species and aquifer sediments are largely influenced by the associated biogeochemistry of the aquifer systems. The amount of dissolved oxygen present in the systems and the rate of consumption of dissolved oxygen generally play an important role in controlling the redox potential. As long as the water table remains close to the ground surface, the dissolved oxygen content in groundwater remains in equilibrium with atmospheric oxygen. Consumption of oxygen by microorganisms can shift the equilibrium from oxic to suboxic states. The amount of atmospheric oxygen diffusion into the groundwater through the unsaturated porous medium basically depends on the depth to the water table and the oxygen diffusion rate (ODR) (Mukhtar et al. 1996). The ODR into

Figure 1. Probability of exceedence of arsenic above threshold level (50 ppb) is higher in the SE (southeast) and SW (southwest) zones of Bangladesh.

174

the groundwater varies inversely with the depth to the water surface of the aquifer. If water table moves downward, supply of atmospheric oxygen will be less in the groundwater. The dissolved oxygen level in aquifer recharge water is attributed to the surface runoff and seepage from river bed. Most recently, upstream diversion or withdrawal of river water from the major river systems (Ganges and Jamuna rivers, Fig. 1) and large scale installation of shallow wells in Bangladesh is believed to be responsible for rapid lowering water table. Rapid consumption of dissolved oxygen in recharging groundwater due to the lowering of water table and microbial use of organic carbon as their energy-supplying-electron donor might have triggered the arsenic contamination problem in the groundwater of Bangladesh. This is the principal working hypothesis of this study. This hypothesis is in complete contrast with the existing oxidation theory of arsenic release but partially supports theory of arsenic release by reduction. The main purpose of this investigation was to examine how the shortage of dissolved oxygen in the recharging water contributed to groundwater arsenic pollution problem in Bangladesh. The second purpose is to understand whether the concept of dissolved oxygen shortage in recharging groundwater does or does not contradict the current theories of arsenic release by oxidation and or reduction. The third purpose of this study is to understand how dissolved oxygen shortage might influence the microbial activities that eventually influence the arsenic speciation processes. Addressing these questions may help in finding a suitable bioremediation solution to groundwater arsenic problem in Bangladesh.

2

METHODOLOGY OF THE STUDY

To accomplish the purposes of this study, five analytical approaches were adapted; (1) statistical analyses of spatial distribution patterns of arsenic concentrations, (2) a multiple layer oxygen diffusion model analysis to estimate the amount of dissolved oxygen concentration in deeper layers, (3) analyses of hydrological factors such as hydraulic gradient, groundwater flow patterns and their relation to arsenic contamination problem and dissolved oxygen shortage, (4) computation of redox potential values at different depths of the aquifer and their correlation with vertical arsenic concentration distribution, and (5) thermodynamic analyses and explanations of existing theories of arsenic release and the validity of the hypothesis presented in this study. Thermodynamic analyses were done to establish a link between dissolved oxygen shortage, lowering of water table and their impact on thermodynamic stability of arsenic species and redox potentials values. 2.1

Statistical analyses of arsenic concentrations distribution patterns

The analyses of spatial arsenic distribution patterns were done to understand the impact of dissolved oxygen shortage on the arsenic release mechanism as a function of the associated hydrogeological conditions. A question is, “Which hydrologic zone in Bangladesh did correlate best to the arsenic concentration and how the dissolved oxygen shortage in that process does increase arsenic concentration?” The arsenic concentration records of about 3500 well water samples were analyzed by the BGS/DPHE. The water quality data were analyzed at different universities in the USA and abroad. The authors computed the probability of arsenic concentrations exceeding the threshold level (50 ppb) by using Gumbel exponential distribution method. Based on the volume of available water resources (rainfall, recharge, and surface water), Bangladesh was divided into six hydrological planning zones (MPO 1987) such as, northeast (NE), north-center (NC), north-west (NW), southeast (SE), south-center (SC) and southwest (SW) zones (Fig. 1). The coordinates (latitude and longitude) of the arsenic contaminated well records were inserted into the map of Bangladesh. The distribution patterns of arsenic concentration were statistically analyzed within the boundary of each zone and results are presented in Table 1. It was found from the analyses that the probability of arsenic exceeding the threshold level (50 ppb) in the south-east (SE) zone was computed as 71.4% by using Gumbel equation. The probability of exceeding arsenic below threshold levels 175

Table 1. Probability of exceedence of arsenic concentrations in different hydrologic planning zones of Bangladesh. Hydrologic zone

No. of wells

Average As conc. (␮g/L)

Northeastern (NE) North center (NC) Northwest (NW) Southeastern (SE) Southwest (SW) South center (SC)

1039 192 1072 295 474 295

34.0 (0.5–572) 68 28.6 (0.5–284.0) 51.4 12.3 (0.5–708) 47.1 174.1 (0.5–1090) 199. 84.8 (0.5–1660) 145 38.6 (0.5–862) 113.18

St. Dev Probability of exceedence (%) 10 ppb 58.7 53.9 44.9 80.1 66.2 53.9

25 ppb 48.7 48.0 32.7 76.2 61.4 48.05

50 ppb 34.0 38.9 18.1 71.4 53.3 38.95

100 ppb 14.9 24.4 5 59.55 38.74 24.42

250 ppb ⬍1 4.9 0.3 29.1 12 4.98

Figure 2. Conceptual oxygen diffusion model to estimate oxygen concentrations at deeper layers before and after lowering of water table.

(50 ppb) in the northwest (NW), north center (NC) zones are shown in Figure 1 and the computed values are tabulated in Table 1. 2.2

Multi layer oxygen diffusion model to estimate the dissolved oxygen level

In order to determine the impact of shortage of dissolved oxygen in the recharge water at deeper layers of aquifers as a function of lowering the water table, a numerical oxygen diffusion model was built using the Finite Element Analyses technique. The Finite Element technique was used because of its higher accuracy and adaptabilities to numerical solutions for physical processes like convection diffusion, pollution distribution and contaminant transportation. 2.2.1 Conceptual oxygen diffusion model For the sake of simplicity, this oxygen diffusion model considers a constant oxygen consumption rate (␣ ⫽ 0.0021 cm3/cm3/hr) by organic carbon in both saturated and unsaturated zone of the aquifer. This model was built to validate the theory of arsenic release by oxidation where the arsenic rich aquifer layers are exposed to the atmospheric oxygen. It was hypothetically assumed that the arsenopyrite-rich sediments layers L-6 an L-10 are located respectively at 6 and 10 meter below the ground surface (Fig. 2). A layered 10 m thick aquifer was modeled, representing the layer numbers by L-1 to L-10 (Fig. 2) with each layer being 100 cm thick. WT-1 and WT-2 show 176

Figure 3. Change in dissolved oxygen concentration in layer L-6 and L-10 after lowering of water table 5 m from the position of WT-1 to WT-2.

the water table elevation before and after large-scale well installation and Case-1 and Case-2 reflect hydrologic conditions respectively before and after large-scale well installation. This model estimated the change in amount of dissolved oxygen at the exposed deeper layers (L-6 and L-10, Fig. 2) after lowering water table from WT-1 to WT-2. The oxygen diffusion model was simulated up to 350 hours from beginning with time step ⌬t is equal to 1 and 4 hours. However, results of 1-hour time step are printed every 10 hours (Fig. 3). The change in oxygen exposure took place at layer L-6 and L-10 before and after introduction of large-scale shallow irrigation wells in Bangladesh was investigated. In the diffusion model the following values were addumed: the diffusion coefficient Dx was 259.2 cm2/hr (4.166 * 10⫺4 m2/s) and the oxygen diffusion rate was ⫺ 0.002125 cm3/cm3/h. The results of the transient oxygen diffusion model were analyzed and plotted (Fig. 3). Figure 3 shows that a 4 m lowering of the water table from layer L-2 to L-6 did not result in an increase of the oxygen concentration at L-10 and L-6. The oxygen concentration in Case-1 at layer L-6 was predicted 0.09 atm (3.64 mg/L) after 150 hours. After lowering of the water table, using the same time interval and the same aquifer properties, the oxygen concentration was found to be 0.06 atm (2.39 mg/L). This result implies that lowering of the water table cannot increase the oxygen supply to the deeper layers of the aquifer. These results are help to establish a link between current theory of arsenic release and role of oxygen shortage and are further addressed in the results and discussion section. 2.3

Hydrological factors and their relation to arsenic contamination problem

The relations among arsenic distribution patterns and groundwater velocity and flow directions were analyzed by dividing the whole country into a number of square grids where arsenic concentration records and groundwater flow direction maps were available. The whole country was divided into 16 177

Figure 4. The schematic groundwater flow direction is shown towards the main river system and the Bay of Bengal where the difference between the mean sea level (MSL) is about 95 m (from north to south). 1200 Arsenic concentration and groundwater flow direction Arsenic concentration(ppb)

1000

800

600

400 Increasing trend of arsenic contamination

200

0 6 7 8 9 Location of grid numbers along groundwater flow direction (North to South)

Figure 5.

10

Increased trend of arsenic concentration from north to south along the groundwater flow direction.

178

Figure 6.

Arsenic concentration distribution over aquifer depth.

grids where arsenic contaminated wells were available. Grids 6, 7, 8, 9 and 10 (Fig. 4) have the same general hydraulic gradient and direction of flow. The arsenic concentration records in those five grids showed a strong correlation between arsenic concentration and groundwater flow direction (Fig. 5). 2.4

Computation of redox potential values and their relation with arsenic concentration distribution

The redox potential values at different layers of the aquifer system were computed by measuring the activity of electrons in a solution expressed in units of volts (Eh) or in units of electron activity (pe). The aquifer sediment contains solid amorphous Fe(OH)3 and it was assumed that the potential corresponds to an oxidation-reduction potential of the aquatic environment. By computing equilibrium constant value of log K, and the standard state free energy value ⌬G0, the equation that expressed the equilibrium of iron oxides with dissolved Fe2⫹ gave the value of the redox potential in the groundwater systems (Figs. 7 and 8).

3

RESULT AND DISCUSSIONS

The results of the analyses are discussed in the light of the hypothesis of this study. The working hypothesis was that layers at or below the water table receive less oxygenated water because of lowering water table which is due to large scale installation of irrigation wells and upstream withdrawal of river water in Bangladesh. The shortage of dissolved oxygen in the recharge water may also be due to microbial activity. The immediate question one would ask is how a shortage of dissolved oxygen occurred in the recharging groundwater and how it contributes to arsenic mobilization. It should be kept in mind that the cause of arsenic contamination in the groundwater of Bangladesh is still poorly understood and a strongly debatable issue. However, the above questions could be answered in the context of hydrogeochemical issues. It was apparent from Table 1 that the probability of arsenic exceeding threshold level (50 ppb) was highest in the southeast (SE) and 179

Redox potential, pe (V)

Aquifer depth (m)

Redox potential values (pe) decrease with aquifer depth from 10 to 30 m below the surface.

Redox potential, pe (V)

Figure 7.

Aquifer depth (m)

Figure 8. The redox potential values (pe) start increasing with aquifer depth from 150 to 350 m below the surface indicating that the recharge water is rich in dissolved oxygen.

southwest (SW) zones of Bangladesh. These two regions were built by the sediment of the Meghna and Ganges River Flood Plain. This delta usually experiences a high rate of sediment flow of about 479 million tons per annum (BWDB 1993). The least arsenic contaminated regions are located in the northwest (NW) and north center (NC) regions where the surface geology is built up by the Old Himalayan Fan, Tista-Jamuna Flood Plains, and Madhupur Tract. The top of the main aquifer systems in the NW zone is closer to the surface than that of the SE and SW regions. The regional groundwater gradient in the NW zone is 2 m/km but the gradient in the SE and zone is 0.1 m/km (MPO, 1987). It is apparent that the groundwater gradient in NW zone is 20 times higher than southern zones(SE & SW) where the arsenic concentration is maximum (Fig. 1). 3.1

Arsenic concentration is greater in the southeast (SE ) and southwest (SW) zones

The sedimentology of northwest (NW) zone is different from SE and SW zone. Moreover, the hydraulic gradient in NW zone is twenty times higher than SW zone; therefore, the residence time of groundwater in NW zone would be less than SW and SE zones. On the contrary, the SW and SE zones were mostly built up through the sedimentary deposition, and the hydraulic gradient in the SW and SE zones is less; therefore, the groundwater residence time is expected to be higher than 180

in the NW zone. Potential recharge water in the shallow aquifer is usually rich in dissolved oxygen. Under the influence of the high hydraulic gradient in NW zone, the oxygenated groundwater recharge can quickly and easily replace the old groundwater. After rejuvenating the aquifers by recharge water, the groundwater in NW zone is expected to remain oxic if there is not too much organic carbon leaching into the aquifer from the top of the ground surface. Continuous percolation of organic carbon from the increased agricultural activities can also result in rapid consumption of dissolved oxygen in the groundwater. The groundwater in maximum arsenic contaminated area (SW and SE zones, Fig. 1) might be continuously lacking in dissolved oxygen because it can not be easily replaced by the oxygenated recharge water because of low gradient. As a result, the redox potential value always remains low in SW and SE region. Only mixing with oxygenated water or lack of reductant minerals can maintain the desired oxic condition in SW and SE regions, but achieving oxic conditions is difficult in these zones due to rapid consumption of dissolved oxygen. Therefore, arsenic mobilization is the higher in SW and SE region than NW, NE and north center (NC) zone in Bangladesh. 3.1.1 Lack of correlation between arsenic concentrations and depth of water table The investigation report (BGS, 2001) showed that arsenic contamination does not have any relationship with depth to the water table or amount of irrigation withdrawal rate in Bangladesh. Moreover, in the most contaminated zone (SE and SW zone) the water table does not match with the most intensive extraction rate. Because both arsenic release mechanisms (oxidation or reduction) are time dependent processes but concentrations spatially differ, it is inappropriate to explain differences using a space–dependent relationship. Therefore, it might be misleading to establish a direct correlation between arsenic concentrations and depth of water table or extraction rate (Q). For example, if the volume of irrigation withdrawal is considered as a function of water table depth, a consistent relationship can be obtained only when the specific yields and effective porosity are the same for every aquifer. Otherwise, the correlation will be influenced by those parameters and may destroy the relationship. The lack of correlation between arsenic concentrations and lowering of water table or abstraction rates (Q) does not necessarily lead to rejecting the idea that over-pumping of irrigation wells may cause arsenic mobilization in the groundwater of Bangladesh. Recently, a group of researchers based in Manchester University found evidence that influxes of organic carbon in groundwater are known to occur when irrigation wells are drawn down. This group also found that introduction of organic carbon by over-pumping of irrigation wells can be a factor in increasing arsenic mobility in shallow groundwater in Bangladesh (Roach 2004). Organic carbon in the sediment acts as a food source of bacteria. Bacteria would consume dissolved oxygen and eventually could lead to change the biogeochemistry of the arsenic contaminated groundwater. 3.2

Maximum arsenic concentration at the depth 30 to 50 meter below the ground surface

The analysis of vertical distribution pattern of arsenic concentrations is shown in Figure 6. That shows that arsenic concentrations above threshold level (⬎50 to 1600 ppb) are mostly confined within between 30 and 50 m below the ground level. But, arsenic concentrations were less than 50 ppb between 150 and 350 m. The sequence of vertical arsenic distribution can be interpreted in terms of redox potential values as shown in Figures 7 and 8. The increasing trend of arsenic at the depth ranging from 9 to 50 m below the ground (Fig. 6) was found to be associated with decreasing trend of redox potential as shown in Figure 7. The decreasing trend of arsenic from 150 to 350 m in Figure 6 is related to the increasing trend in redox values as shown in Figure 8. Although, the two aquifer systems have many differences, the sequence of decreasing and increasing trend of the redox potential values (Figs. 7 and 8) could be associated with the dissolved oxygen content in the groundwater. The recharging groundwater at the depth of 150 to 350 m is mostly attributed to the regional thorough-flow which is rich in dissolved oxygen. The presence of electron donors at the deeper aquifer layers is also less. On the other hand, the recharging groundwater at the upper part of the aquifer is also rich in dissolved oxygen since it comes from the surface runoff and river 181

bed seepage, but this dissolved oxygen content can be quickly consumed by the electron donors. Therefore, the shortage of dissolved oxygen might play an important role in mobilizing arsenic in the groundwater system because dissolution of iron oxides occurs under anoxic conditions (Lee et al. 2003). 3.2.1 How does the dissolved oxygen level in recharging water decrease over aquifer depth? It is a common misconception that lowering of water table would increase the unsaturated vadose zone and eventually the oxygen concentration level in deeper layers of the aquifer will be increased. The oxygen diffusion model analyses as shown in Figure 2 demonstrated that lowering of water table does not increase the oxygen supply rate into the deeper layers of the aquifer. To prove this, a one dimensional oxygen diffusion model was built, where the water table elevations at WT-1 and WT-2 (Fig. 2) represent the levels before and after installation of irrigation wells respectively. The oxygen diffusion model suggests that lowering the water table about four meters (from WT-1 to WT-2) did not increase the dissolved oxygen concentration at the hypothetical arsenic contaminated layers at L-6 and L-10 (Fig. 2). The oxygen concentration in Case-1 at layer L-6 was estimated as 0.09 atm (3.64 mg/L) after 150 hours. But in Case-2, after lowering of the water table, using the same interval of time and same aquifer properties, the oxygen concentration at L-6 was found as 0.06 atm (2.39 mg/L). Therefore, lowering water table decreased the dissolved oxygen concentrations in the same aquifer. Consumption or reduction of dissolved oxygen supply helps develop reducing conditions. 3.2.2 Shortage of dissolved oxygen decreases the redox potential values The vertical arsenic distribution patterns as shown in Figure 6 indicate that the wells contaminated with arsenic in concentrations greater than 50 ppb are mostly located at the depth ranging from 10 to 35 m below the ground. The associated redox potential values (pe) were computed using the dissolved iron concentration records and are depicted in the Figures 7 and 8. It is apparent in Figure 7 that the redox potential values start decreasing (0 to ⫺0.5) from the surface of the water table and continue up to 35 m below the ground level where the pe values fall below zero (⫺1.5) (Fig. 7). Certainly, the main reason for the decreasing redox value is the consumption of dissolved oxygen by electron donors. But the redox potential values at the top most surface of the water table are higher than the deeper part of the aquifer because of its close proximity to the ground surface. At the top layer of the aquifer, the dissolved oxygen level is detectable and much higher than deeper layers. The redox potential values again start increasing from 150 to 350 m below the ground. This is also associated with the dissolved oxygen concentration in the recharging groundwater. This can be reasoned that recharging groundwater at the depth below 150 m is known as the regional throughflow. The regional through-flow is usually rich in dissolved oxygen level, and in the deeper layers the oxygen consumption rate is negligible since organic carbon can not reach that point. 3.3 Validity of the existing theories of arsenic release from the context of dissolved oxygen shortage There have been divergences of views among the researchers and considerable difficulties in explaining the cause of arsenic mobilization in Bangladesh groundwater. While it was not the purpose of this analysis to focus on this debate, it was necessary to evaluate both oxidation and reduction theory of arsenic release in the context of shortage of dissolved oxygen in order to support the working hypothesis. Since, the working hypothesis of this study directly opposes the oxidation theory of arsenic release; inadequacies in the oxidation theory will be discussed. The proponents of the reduction theory have rejected the explanations of oxidation theory of arsenic release. They argue that if pyrite oxidation were the real cause of arsenic release, arsenic concentrations must have been positively correlated with the amount of sulfate in the arsenic contaminated groundwater. Arsenic contaminated groundwater in Bangladesh does not provide any evidence of correlation between arsenic and sulfate in the field. However, Bridge & Hussain (2000) mentioned that the lack of correlation between arsenic and sulfate does not indicate the 182

rejection or acceptance of pyrite oxidation theory. They reasoned that sulfate concentration in groundwater depends on many factors like pyrite grain size, pyrite abundance, reaction rate, migration time, etc. In order to examine the oxidation theory of arsenic release, the authors argue that there must be arsenic concentration gradients during pumping from the contaminated shallow wells. If oxygen exposure due to lowering of water table were the valid reason of arsenic release, then maximum arsenic concentration should be located at the topmost layer of the aquifer (0 to 15 m) because dissolved oxygen is the highest in the topmost layer. Since it takes time for arsenic contaminated water to reach the screen of the pumping wells a concentration gradient is expected to occur during pumping. But arsenic contaminated wells do not show any concentration gradient during pumping in Bangladesh. In addition, the oxidation theory of arsenic release can not explain why arsenic concentration is maximum at 30 m below the ground level and why contamination is not maximum just at the top of the water table. Therefore, oxidation theory is inadequate to explain the arsenic mobilization in Bangladesh groundwater. The iron oxides reduction model of arsenic release is now widely accepted as valid for the Bengal Basin (Anwar et al. 2003, McArthur et al. 2001, Ravenscroft et al. 2001), however, their suggestion that reduction is driven by buried peat has not been accepted by all. (Harvey et al. 2002) suggest that reduction of FeOOH is driven by surface organic matter from river beds, ponds, and soils that is drawn into the aquifer by irrigation drawdown. Another unanswered question to support the reduction theory is that if organic carbon is the major cause of arsenic release, why was arsenic contamination not found in the earlier days, and why is there not any prior report of arsenic toxicity in Bangladesh (Adel 2000, Bridge & Hussein 1999). 3.4

Role of dissolved oxygen to influence the microbial activities in arsenic immobilization process

Immobilization of toxic metals and radio nuclides are usually brought about by precipitation, biosorption and bioaccumulation processes. Immobilization by co-precipitation of arsenic with ferric oxides totally depends on amount of dissolved oxygen present or oxic condition. Therefore, dissolved oxygen plays an important role in immobilizing arsenic onto the surface of iron oxides. Biosorption of toxic metals and radionuclide is based on non-enzymatic processes such as adsorption. Adsorption is due to the non-specific binding of ionic species to cell surface-associated or extra cellular polysaccharides and proteins (Gazso 2001).Therefore, the dissolved oxygen content in recharging groundwater may influence the biosorption process of arsenic removal. 3.4.1 Groundwater flow direction and its relation to arsenic mobilization The groundwater flow direction maps (Figs 4 and 5) demonstrate a general trend that arsenic concentrations increased as the groundwater moved from the recharging (NW zone) to the discharging southeast (SE) zone. The groundwater residence time increases with decreasing hydraulic gradient in the southeast zone of Bangladesh. Over time, the dissolved oxygen is consumed up by the electron donors and helps develop a mild reducing condition. Usually, the tendency to donate electrons is high in reducing conditions. Metal-reducing bacteria “breathe” by passing electrons onto metals such as iron to get energy from their food. Iron reducing bacteria may use ferric oxides as energy-supplying electron donors in the groundwater of Bangladesh (Roach 2004). Therefore, arsenic from the surface of iron oxides is released as a result of terminal electron accepting process. Also, the groundwater of SW and SE regions were found highly anoxic ( no trace amount of DO) and offered a strong correlation with dissolved iron content (Nickson et al. 2000). Highly anoxic groundwater in the SE zone of Bangladesh might be attributed to the lower hydraulic gradient and higher residence time than the NW zone. This is why arsenic concentration in the SW and SE zone is higher than the NW hydrologic zones in Bangladesh. A groundwater flow model analysis by the authors (Islam & Bernuth 2003) demonstrated that in general, pumping conditions provide more flow through the deeper layers than the natural groundwater flow conditions. When the pumping rate is greater than the shallow aquifer recharge (SAR) 183

rate, it increases the flow through the deeper layers when the aquifer is connected to the river system. On the contrary, when SAR is greater than the pumping rate, it reduces the flow through the deeper layers. The volume of SAR water is partially attributed to the river water and the river water is usually rich in dissolved oxygen. Therefore upstream withdrawal of river water and the higher extraction rate of wells would greatly reduce the SAR rate. Under that situation, groundwater flow rate through the deeper layers would be increased (Islam & Bernuth 2003) and may contribute to arsenic mobilization in Bangladesh groundwater.

4

CONCLUSIONS

Groundwater in the aquifers close to the surface is usually rich in dissolved oxygen (DO). The presence of dissolved oxygen in the recharging groundwater is mainly attributed to the seepage from river beds, close proximity of water table to the surface, surface runoff and lack of organic carbon present in the system. Microbial growth influences the DO concentration in the aquifer. It is apparent that the arsenic contamination in Bangladesh is not directly relational with depth to the water table or extraction rate. Analyses have shown that lowering the water table and extraction rate may influence the arsenic mobilization but that it is further influenced by the dissolved oxygen content. Recent studies have demonstrated that lowering the water table and/or the extraction rate influence the amount of organic carbon available in the deeper layers and that is believed to be responsible for arsenic mobilization. In addition, the pumping rate has a significant influence on the flow patterns through the deeper layers of the aquifer system. The groundwater in the deeper layers is expected to be anoxic since the recharging water may come from the storage if extraction rate is higher than aquifer recharge (SAR) rate. Continuous pumping at a rate greater than SAR may lead to lowering of water table which will eventually reduce the dissolved oxygen supply to the groundwater and would contribute to the arsenic mobilization in Bangladesh groundwater. Since arsenic mobilization is not solely due to chemical changes; but rather it is a resultant of complex multidimensional biogeochemical and hydro-geological processes. Although it is true that arsenic mobilization is poorly understood, but it is most likely that arsenic contamination problem in Bangladesh is attributed to the shortage of dissolved oxygen level in the recharging water. The exact contribution of river water diversion to the shortage of dissolved oxygen in recharging groundwater is to be estimated by reliable and large scale hydrologic model analyses.

REFERENCES Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A., Imam, M.B., Khan, A.A., & Sracek, O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh: An overview. In: Bhattacharya, P., Welch, A.H., Ahmed, K.M., Jacks, G. & Naidu, R. (eds) Arsenic in Groundwater of Sedimentary Aquifers Appl. Geochem. 19(2): 181–200. Adel, M.M. 2000. “Arsenification: Searching for alternative theory”. Daily Star of Bangladesh. 28 April, 2000. Anawar, H.M., Akai, J., Komaki, K., Terao, H., Yoshioka, T., Ishizuka, T., Safiullah, S. & Kato, K. 2003. Geochemical occurrence of arsenic in groundwater of Bangladesh: source and mobilization processes. J. Geochem. Explor. 77: 109–131. Bhattacharya, P., Frisbie, S.H., Smith, E., Naidu, R., Jacks, G. & Sarkar, B. 2002a. Arsenic in the Environment: A Global Perspective. In: Sarkar B (ed) Handbook of Heavy Metals in the Environment (Marcell Dekker Inc., New York): 145–215. Bhattacharya, P., Jacks, G., Ahmed, K.M., Khan, A.A. & Routh, J. 2002b. Arsenic in groundwater of the Bengal Delta Plain aquifers in Bangladesh. Bull. Env. Cont. Toxicology 69: 538–545. BGS (British Geological Survey) & DPHE 2001. Arsenic contamination of groundwater in Bangladesh. Technical report WC/00/19, Volume.2. BWDB. 1993. Hydrological map of Bangladesh. Scale 1:1, 000000. Bangladesh Water Development Board, Motejheel CA, Dhaka.

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Bridge, T.E. & Hussain, M.T. 1999. The increased draw down and recharge in groundwater aquifer and their relationships to the arsenic problem in Bangladesh. URL: http://www.dainichi-consul.co.jp./english/ arsenic/arsarticles.htm. Gazsó, Lajos, G. 2001. The key microbial processes in the removal of toxic metals and radionuclides from the environment. CEJOEM 7 (3–4): 178–185. Harvey, C.F., Swartz, C.H., Badruzzaman, A.B.M., Keon-Blute, N., Yu, W., Ali, M.A., Jay, J., Beckie, R., Niedan, V., Brabander, D., Oates, P.M., Ashfaque, K.N., Islam, S., Hemond, H.F. & Ahmed, M.F. 2002. Arsenic mobility and groundwater extraction in Bangladesh. Science 298: 1602–1606. Islam, N. & Bernuth, R.D. 2003. Arsenic transportation in the groundwater of Bangladesh and the policy of irrigation development. Journal de Physique IV France 107: 661–665. Khan, A.H., Rasul, S.B., Munir, A.K.M., Habibuddowla, M., Alauddin, M. Newaz, S. S. & A. Hussam. 2000. Appraisal of a simple arsenic removal method for groundwater of Bangladesh. J. Environ. Sci. Health A35 (7): 1021–1041. Lee,-Y., Um,-I-H & Yoon, J. 2003. Arsenic (III) oxidation by iron (VI) (ferrate) and subsequent removal of arsenic (V) by iron (III) coagulation. Environ. Sci. Technol. 37(24): 5750–5756. McArthur, J.M., Ravenscroft, P., Safiullah, S. & Thirlwall, M.F. 2001. Arsenic in groundwater: testing pollution mechanism for sedimentary aquifers in Bangladesh. Water Resour. Res. 37: 109–117. Mukter, S., Baker, J.L. & Kanwar, R.S. 1996. Effect of south term flooding and drainage on soil oxygenation. Transaction of the ASAE 39(3): 915–920. MPO (Mastar Plan Organization). 1987. The groundwater resources and its availability for development. June 1987. Technical Report No. 5. Submitted to the Ministry of Irrigation, Water Development and Flood Control, Government of the Peoples Republic of Bangladesh assisted by the United Nations Development Program (UNDP) and the World Bank. Megonigal, P., Emerson, D. & Weiss, J. 1997. Iron-oxidizing bacteria occur in the rhizosphere of wetland plants at circumneutral pH. BIOGEOMON ’97. Journal of Conference Abstracts 2(2): 245. Nickson, R.T., McArthur, J.M., Ravenscorft, Burgess, W.G. & Ahmed K.M. 2000. Mechanism of arsenic release to groundwater of Bangladesh and West Bengal. Appl. Geochem. 15: 403–413. Roach, J.. 2004. Arsenic in Asian Drinking Water Linked to Microbes. Published in National Geographic News. URL: http://news.nationalgeographic.com/news/2004/06/0630_040630_arsenic.html#main. Ravenscroft, P., McArthur, J.M. & Hoque, B.A. 2001. Geochemical and palaeohydrological controls on pollution of groundwater by arsenic. In: Chappell, W.R., Abernathy, C.O., Calderon, R.L. (Eds.), Arsenic Exposure and Health Effects IV pp. 53–77. Elsevier, Oxford. Stumm, W. & Morgan, J.J. 1996. Aquatic Chemistry, Chemical Equilibrium and rates in natural water, 3rd Edition, Wiley Interscience.

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Section 4: Remediation of arsenic-rich groundwaters

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Technologies for arsenic removal from potable water W. Driehaus GEH Wasserchemie GmbH & Co. KG, Osnabrueck, Germany

ABSTRACT: Arsenic (As) contamination of ground water resources used for potable water supply is an emerging problem throughout the world. Except for the use of alternative, uncontaminated water sources, water treatment for As removal is often the only solution to meet the standards. The effective application of As removal processes requires the knowledge of its chemistry in natural water. Before making a choice of a suitable treatment technique it is to ascertain whether As(III) is present and to evaluate the need for an oxidation technique for As(III). The techniques existing for the removal of As include conventional processes like ion exchange and coagulation/filtration and also emerging processes with iron oxide based adsorbents like granular ferric hydroxides. While most of the removal techniques suitable for water works operation rely on sorption processes, they exhibit big differences in investment, chemical cost, maintenance and the type of arsenic bearing waste. Different options for disposal are discussed.

1

INTRODUCTION

Arsenic (As) became a serious health problem in several countries around the world. This is also attributed to new findings about the toxicity of arsenic, especially for long term exposure to low levels in drinking water and food. Thus, a number of countries have lowered their drinking water standards within the last 13 years (WHO 2001). The lowered standards, mostly to 10 ␮g/L and the “As calamity” in India and Bangladesh induced world wide activities in research of available treatment methods for As-removal and the development of new technologies. The situation of drinking water supply in the affected countries and areas are different, from more centralized treatment plants and supply i.e. in the UK and the USA, to smaller decentralized waterworks in Germany and France to a supply by hand pumps, missing any distribution network in the affected areas of West Bengal and Bangladesh. This paper gives an introduction to the basics of aquatic chemistry of As and an overview on As removal techniques in potable water treatment, including oxidation techniques for As(III).

2

AQUATIC CHEMISTRY OF ARSENIC

Arsenic is a semimetal and occurs in natural groundwater mostly occurring as oxyanions as trivalent As (H3AsO3) or pentavalent As (H3AsO4). They are commonly named arsenite or As(III) and arsenate or As(V). The occurrence of organic As compounds, especially methylated species is only reported from surface water (Anderson & Bruland 1991), but they rarely, if ever occur in groundwater. Elevated As concentrations at a level relevant for human health are mostly caused by inorganic As species in groundwater. Figure 1 shows the stability diagram of inorganic As. Natural groundwater is commonly in the field between pH 5–9 and Eh ⫺0.50 V to ⫹0.50 V. Under this conditions there are two dominating species, As(III) in more reducing environments and As(V) under oxidizing conditions. In strongly reducing environments with hydrogen sulfide present form also a couple of As-sulfides, most of them being solids. As this conditions are rare in 189

Figure 1.

Stability diagram for inorganic As (according to Ferguson & Gavis 1972, Baldauf 1995).

groundwater bodies used for drinking water supply, these species are excluded from the following description. The dissociation constants of As(III) are pKS1 ⫽ 9.22; pKS2 ⫽ 12.10; pKS3 ⫽ 13.40 (Pierce 1981). That means that at pH 9.2 the As(III) is by 50% dissociated. At lower pH, most of As(III) exists as a neutral molecule. The dissociation constants of As(V) are pKS1 ⫽ 2.22; pKS2 ⫽ 6.96; pKS3 ⫽ 11.5. That means, that a pH 6.96 about 50% of As(V) exists as a monovalent anion and 50% as an divalent anion. The difference in charge has important effects on the removal characteristics of both As species, because neutral, uncharged molecules can not or less effectively be removed by most treatment techniques (Jekel 1994). As(III) can be oxidized to As(V) at a relatively low Eh-potential of 0.1–0.2 V and, from the energetic point of view, dissolved oxygen is sufficient as an oxidant. Unfortunately, the oxidation of As(III) by oxygen is very slow with conversion rates of a few percent per day. Thus, even oxidizing groundwater with high oxygen concentrations may contain some As(III). As(III) in its typical stability range is very often associated with dissolved iron and manganese. As(V) is the stable species in oxidizing water and has a great similarity to ortho-phosphate in terms of dissociation, precipitation, adsorption and ion exchange (Jekel 1994). Thus phosphate is identified as one of the major competing ions in As(V) adsorption.

3

TREATMENT TECHNIQUES

Nearly all existing treatment technologies have been examined for their ability to remove As from water. In addition, new techniques were developed to avoid the shorts of existing techniques. The principle goals for As removal techniques are:

• • • •

Efficiency: output concentrations below the standard of 10 ppb should be safely achieved, Reliability and maintenance: Techniques should be reliable and adapted to the skills of the staff. Residuals: Take into account the handling, treatment (if necessary) and disposal of any treatment residuals, they are suspected to contain high amounts of As. Costs: Investment as well as operating costs define whether the treatment technique is economical. 190

Figure 2.

Classification of As-removal techniques.

The more promising treatment techniques for As removal, which are currently applied, are (USEPA 2000): (I) Technologies based on adsorption reactions: • Coagulation with iron or alum coagulants • Iron and manganese removal • Subterranean removal • Adsorption on activated alumina • Adsorption on iron oxide based adsorbants (II) Other technologies: • Ion exchange • Membrane processes, i.e. reverse osmosis and nanofiltration. Figure 2 gives schematically an overview for the classification of As-removal processes. We distinguish between the sorption processes and the separation processes. Sorption processes means here, that chemically active solids are involved in the process and that the contaminant is bound by electrostatic or chemical forces to the surface. Ion exchange generally requires a charged ion and thus, it fails completely in removing As(III) at neutral pH. The adsorption techniques are all based on the great affinity of As species to metal oxide surfaces, to form stable surface complexes (chemisorption). Some surfaces have the potential to bind As(III), but mostly to a minor extend compared to As(V). Arsenic removal by adsorption on metal oxides and hydroxides is the most important removal mechanism in natural environments (Soils, lacustrine and marine sediments) and also in technical applications. The adsorption reaction is understood as a surface complexation reaction, where As forms inner-sphere complexes. Metal oxide surfaces are usually protonated and contain positive and negative electrostatic charges at the different surface sites, according to the following reactions (M stands for metals like Fe or Al). 191

Protonation and deprotonation of metal oxides:

Adsorption of As(V) on metal oxide surfaces:

Protonation and deprotonation reactions are pH dependent and at a distinct pH, the isoelectric point, the surface contains negative and positive charges to the same extend and the overall charge is zero. The pH of the isoelectric point varies for different metal oxides. Iron(III) hydroxides and oxihydroxides have their isoelectric point generally around pH 7–8. At greater pH, the overall surface charge is negative, below, the overall charge is positive. This has a big impact on the adsorption of As(V): At neutral pH, As(V) as an anion is negatively charged. It will adsorb very good to a positively charged surface, but the adsorption is hindered by a negatively charged surface. The specific adsorption of As takes place via an oxygen bridge between the metal atom and As. The reaction given above shows a monodentate complex, but also bidentate complexes occur, which are using two sites for one As atom (Waychunas et al. 1993). At high pH adsorption and removal efficiency is reduced and vice versa. Thus, removal techniques for As(V), which rely on adsorption work far better at low pH. In the practical consequence one need to dose more coagulant for the same removal result, or, with fixed bed techniques, one need to change out or regenerate the sorbent more frequently. The membrane processes are separation processes and remove nearly all dissolved substances from the water. They are also uneven in the removal of As(III) or As(V). Usually, As(V) is much better removed than As(III). Uncharged molecules have, in any way, the ability to pass membranes to some extend. 3.1

Oxidation techniques for As(III)

Most of the treatment techniques are much more effective in removing As(V) rather than As(III), because they partially rely on electrostatic forces. Thus, there is a great discussion on oxidation techniques for As(III). With As(III) present in a raw water, usually other reduced species are present: ferrous iron and manganese(II). This co-occurrence is important for the evaluation of oxidation processes because iron and manganese are also subject to treatment goals. As described previously, As(III) can not be oxidized simply by dissolved oxygen, because this reactions are kinetically hindered and very slow. In this way, stronger oxidants have to be applied. The following oxidants are applied successfully for As(III) oxidation (Jekel 1994):

• • • • •

Chlorine, hypochlorite Ozone Permanganate Hydrogen peroxide Manganese oxide

Chlorination is one method for As(III) oxidation, either by dosing gaseous Cl2 or by dosing a hypochlorite solution. The required dose is between 0.3–1 mg/L of Cl. The dose depends mainly on the Cl-consumption by other reduced species like ferrous iron, manganese(II) and organics. Restrictions and concerns occur about the formation of chlorination byproducts. They evolve by the reaction of natural organic matter (NOM) with chlorine and are generally called THM ⫽ Trihalomethanes. The oxidation of As(III) by chlorine is a fast and complete reaction. This and the fact, that chlorination is a very common chemical in the preparation and treatment of potable waters, makes it to the most commonly applied oxidation method for As(III). 192

Ozone (O3) is another potent oxidant, effective for As(III) oxidation. Same as Chlorine, ozone oxidizes also other reduced species, like ferrous iron and manganese(II). The reaction mechanism is via hydroxyl radicals. Despite the different mechanism, there are similar concerns about the side effects as with chlorine, the oxidation byproducts. This byproducts are produced during partial oxidation of NOM. Ozone has to be prepared on site by a high voltage discharge process and cannot be stored in tanks. Ozone is applied in water treatment processes for nearly all oxidation processes, also those to correct taste deteriorations by oxidizing organic compounds. The oxidants permanganate and hydrogen peroxide are used in a low number of treatment facilities to oxidize As(III). Especially permanganate has a good potential to be used for As(III) oxidation. It reacts fast, leads to an almost complete conversion and the As(III) oxidation is preferred over other specific oxidation processes like ferrous to ferric iron (Borho 1996). The application of permanganate requires a filtration step to remove the manganese oxide particles which are generated during its reaction, that can simultaneously be used to remove the oxidation product arsenate(V) after dosing of coagulants. A single hydrogen peroxide application cannot oxidize As(III) quick enough. If ferrous iron is present in the raw water, hydrogen peroxide reacts with Fe2⫹, which is known as the Fenton’s reaction. The Fenton’s reaction generates hydroxyl radicals, which act as oxidants for As. Since some iron hydroxide is also generated by this reaction, As(III) can be oxidized and partially adsorbed to the suspended ferric hydroxide in one step. Same as with Ozone, the Fenton’s reaction produces undesired byproducts by partially oxidizing NOM. The need for an As(III) oxidation has to be carefully evaluated for several reasons:

• • • •

The application of oxidants has sometimes undesired side effects. Arsenic occurs more frequently in its oxidized form as As(V) and an oxidation is completely useless with respect to As removal performance. Some of recently developed removal techniques are also effective in removing As(III). Oxidation by dosing chemicals makes the process more complex. This may be unacceptable, when using maintenance-free fixed bed processes.

Oxidation reactions of As(III) occur also as a side effect in conventional treatment for iron and manganese removal: As(III) is oxidized together with iron and manganese during filtration and removed by adsorption on solid ferric hydroxide. This mechanism is reported from a number of conventional treatment plants in Germany (Haase, personal communication), which contain high As amounts in the residual sludge. There is the evidence of a biological aided oxidation during the oxidation and removal of ferrous iron and/or manganese(II) (Seith & Jekel 1997, Driehaus et al. 1995). One promising technique for As(III) oxidation is also the filtration over manganese oxide media: That can be manganese oxide coated sand (byproduct from groundwater treatment processes) or commercially available mined manganese oxide products. For the recently developed As treatment by adsorption on iron based adsorbents, it is reported that they are also effective for removing As(III), thus eventually avoiding the need of a special oxidation treatment (Driehaus et al. 1998).

4 4.1

ARSENIC REMOVAL TECHNIQUES Conventional techniques

4.1.1 Coagulation/filtration Coagulation, followed by filtration is a widely used treatment, not only for As, but especially for the removal of colloidal substances. There exist also the terms “flocculation” and “precipitation” for this technique, but all mean the same. The coagulation or precipitation technique relies on the dosing of alume or iron salts to form precipitates of iron hydroxide which takes up arsenate by adsorption. The coagulation is applied in combination with a separation technique to remove As loaded particles from the water, i.e. by filtration over granular media, or microfiltration. The 193

residual of this technique is a backwash sludge with a high water content. The sludge needs further treatment for dewatering for easy shipping and disposal. The coagulants of choice are iron salts like ferric chloride (FeCl3), ferric sulphate (Fe2(SO4)3), ferrous chloride (FeCl2) or ferrous sulphate (FeSO4). The use of alum based coagulants is not recommended, because they are far less effective in removing As (Jekel 1994). A scheme of a coagulation/filtration plant for As removal is given in Figure 3. This plant has installed four filter vessels, with two vessels used as a two layer filter for coagulation/filtration. Jekel & Seith (1999) reported a coagulation process at this site being effective for a groundwater source with about 20 ␮g/L As. Arsenic was nearly completely in the oxidized form and no oxidation process was applied. The performance of As removal as a function of the coagulant dose is displayed in Figure 4 (a) for a raw water with about 50% As(III) at a pH of 6–6.2. The removal of As(III) without oxidation is, even at high coagulant doses, very low. Fig 4 (b) shows the percentage

140

As conc. Without chlorine

120

As conc with 1 ppm chlorine

100 pH: 6.0–6.2

80 60 40 20

(a)

0 0

1

2 3 Ferric iron dose, mg/L

4

5

percent As removal

Arsenic conc., µg/L

Figure 3. Scheme of the full scale As removal plant in Germany, where coagulation/filtration and adsorption techniques were examined side by side. (acc. Jekel & Seith 1999).

100 90 80 70 60 50 40 30 20 10 0

control value 5 ppb pH: 7.9–8.0

(b) 0

1

2

3

Ferric iron dose, mg/L

Figure 4. (a) Arsenic removal depends on coagulant dose and oxidation (after Jekel 1994); (b) Iron dose and percent removal at the plant shown in Figure 3 (acc. Jekel & Seith 1999).

194

of As removal versus iron dose from the full scale pilot plant illustrated in Figure 3. An acceptable removal of 85% was achieved with a dose of 1 mg/L as Fe. The coagulant dose depends on the raw water profile, especially on the pH and on concentrations of adsorption competitors like phosphate and silica. For ferric salts, the minimum dose can be calculated from the As concentration. At low pH between 6 and 7 a minimum dose of ferric iron of 10–20 times of the As concentration is required. At higher pH ⭓ 8 the minimum dose is 40–50 times the As concentration. This is only a rough estimation, and it is recommended to test the required coagulant dose by jar tests in the laboratory. The filter(s) separate the As bearing coagulant flocs from the liquid stream. Filters have to be backwashed frequently to remove particles and to avoid a breakthrough of As-loaded particles. The backwash water can be stored in a tank to allow the particles to settle down. After sedimentation the supernatant solution can be discharged or fed again in the process. The main critical step for the implementation and optimization of the conventional coagulation/filtration treatment is the separation of the As loaded particles from the liquid stream. The filtration properties of the particles depend on the chosen coagulant, the dose and the raw water parameters like pH, hardness, alkalinity. The residual from the coagulation process is an As bearing sludge with a high water content. This residual needs a special treatment, at least dewatering, to be handled and transported safely. 4.1.2 Iron and manganese removal, subterranean removal Arsenic removal during iron and manganese removal can be understood as a variation of a coagulation process, because the raw water contains soluble iron, which is oxidized by aeration or by chemical means and precipitated. This technique also needs a filtration step to remove the precipitates from the liquid; mostly applied is the filtration over granular media. There exists the same dependence with regard to pH and competing ions. In general, a sufficient As removal is achieved at a mass ratio of iron to As above 50:1. Some complication arise from the fact, that ground waters with elevated iron concentrations mostly have reduced As(III). This leads to reduced efficiency of the As removal, if it can not be oxidized during treatment. Subterranean removal is a special case of iron removal. The iron containing raw water is aerated and reinjected to the ground, where iron precipitates, adsorbs As and is filtered of by the porous aquifer. The subterranean removal is rarely applied, because it needs special aquifer conditions and it is thought to influence and block the aquifer’s porous structure. 4.1.3 Ion exchange and membrane techniques Ion exchange (IX) and the membrane techniques are less important for water works operation and the application is restricted to special conditions like household treatment in Point-Of-Use and Point-Of-Entry devices. IX is applied as a simple filter technique and needs to be regenerated frequently, the capacity is restricted and other (an)-ions are also removed. Ion exchange (IX), especially anion exchange can effectively remove As, generally using strong base anion exchange resins in the chloride form. However, sulfate and other anions compete with As and can greatly reduce run length. Regeneration is inevitable for ion exchange systems. In natural water, the runlength of IX columns are only 300–1000 bed volumes. Pilot testing indicates that the brine regeneration solution could be reused as about 20 times with no impact on As removal provided that some salt was added to the solution to provide adequate chloride levels for regeneration. This mode of operation would reduce the amount of waste for disposal, but increases the ultimate As concentration of the spent brine. Disposal routes for the spent brine become critical in assessment of the viability of IX for specific treatment situations. The brine can be treated with ferric coagulant to remove the As from the liquid waste. The additional complexity introduced by brine handling and disposal may make IX unattractive for small systems. IX is widely applied and accepted in the USA for potable water treatment. Thus it seems quite attractive to use this technique also for As removal, despite of its shorts as restricted capacity and generation of liquid wastes. There are also a couple of efforts to make improvements to this process. Kim et al (2003) suggested an IX process with a complete brine recycling and treatment, which might 195

overcome the shorts with liquid waste handlings, but could not improve the run length of less than 100 bed volumes. Membrane techniques are not selective to As and other toxic contaminants, but remove most of the dissolved substances and reduce the TDS and conductivity drastically. The membrane techniques remove As(III) to a minor extend and produce a constant stream of waste water, where all removed substances are concentrated. Reverse osmosis (RO) and nanofiltration (NF) provide effective removal of As(V) and nearly all other dissolved species. These technologies are interesting options for point of use and point of entry applications at low flowrates, particularly when As is just one of several water quality parameters requiring treatment. For larger flows, the 15% to 30% feed flow lost as reject (concentrate waste liquid) is an important consideration, as is the 50–150 psi operating pressure required even for modern high efficiency low pressure RO operating on low total dissolved solids feed water. For sites with several water quality issues to address, particularly if these are related to dissolved solids, reverse osmosis or nanofiltration have the attraction of providing complete treatment in a single process step. 4.1.4 Activated alumina The adsorption techniques in the stricter sense rely on a simple filtration process over granular adsorbents like activated alumina or granular iron oxides and ferric hydroxides. Dosing of chemicals is usually not required. These techniques do not produce a backwash sludge, the residual is the As loaded adsorbent itself. Activated alumina is known since more than 20 years as a good adsorbent for arsenate containing waters, but needs a regeneration, due to the restricted lifetime of the media until exhaustion. It is reported that the optimal pH for As removal with activated alumina is around 6.0. Thus a pH-adjustment by adding mineral acids or CO2 could increase the treatment capacity drastically. Figure 5 shows the treatment capacities of activated alumina with a model raw water containing As(III) or, after oxidation, As(V). Obviously is activated alumina not effective in adsorbing As(III). Activated alumina can be regenerated by rinsing with diluted caustic soda at a concentration of 2–4%, followed by rinse with 2% sulfuric to equilibrate to neutral pH. This causes some problems because it produces a significant loss in capacity and, more important, liquid waste streams highly enriched with As. Recently, various activated aluminas were pilot tested with ground water from Arizona. They exhibited treatment capacities between 1000 and 4000 bed volumes at raw water pH of 7.5–9 (Chang et al., in press). The advantages of activated alumina, simple filter operation, and the shorts, restricted capacity, lead to the development of iron oxide based adsorbents in granular form, as it was expected that they have a much higher capacity, while being suitable for fixed bed operation. 4.2

Emerging techniques: Iron oxide based adsorbents

Effluent arsenic, µg/L

The technique with iron oxide based adsorbents was in 1991–1994 developed at the Technical University of Berlin, Department of Water Quality Control, to meet the new treatment goals of the 120 100

As(III)

80

As(V)

60 40 20 0 0

5000

10000 15000 20000 25000 30000 35000 Bed volumes

Figure 5. Process life of activated alumina with 100 ␮g/L As(III) and As(V) at pH 6 (Frank & Clifford 1986).

196

lowered As standard in Germany (Driehaus et al. 1998). The technique provides a simple filtration process over granular adsorbent medias, commonly without any dose of chemicals and without pH adjustment. The following description focuses on GEH, one of several available types of iron oxide based adsorbents. Other iron oxide based adsorbents are currently under development and examination (Zeng 2002). GEH is a pure ferric hydroxide in granular form. The grain size is 0.32–2 mm and the specific surface is 250–300 m2/g. The maximum As adsorption density is 55 g/kg, the typical adsorption from drinking water applications is 1–10 g/kg. The adsorption densities, that are calculated from batch tests at different pH values are given in Figure 6. The adsorption density at a residual concentration of 10 ␮g/L is plotted against pH for both As(III) and As(V). At low pH, the adsorption density of As(V) is much higher than of As (III), but at slightly alkaline pH, adsorption is nearly equal for both oxidation states of As. The adsorption density and the lifetime until exhaustion in a treatment plant increases with decreasing pH for As(V). It is quite constant for As(III). The batch tests were prepared simulating a typical groundwater chemistry with an electrical conductivity of 480 ␮S/cm. Natural water has some constituents which interact with the ferric hydroxide surface and lead more or less to a reduction in adsorption density for As. The most important interfering substances are phosphate, dissolved organic matter and at low pH. Also high silica amounts may interfere with As adsorption, reducing the adsorption kinetics and the treatment capacities (Waltham & Eick 2002). For practical evaluations, for the comparison of different media and for economic calculations, the treatment capacity, expressed in bed volumes, rather than the adsorption density is a useful expression. Figure 7 shows a typical scheme of an installation and an installation in a 32 year old building. Currently more than 80 plants with granular ferric hydroxide are installed worldwide. We give an overview on operational data for 24 plants in EU states. The design flow rates of the yet installed treatment plants are between 4 and 800 m3/h and the annual supply is from 10,000 m3 to 7,000,000 m3. All plants usually work in a non-continuous mode, i.e. with supply by night to a reservoir. The empty bed contact time (EBCT) can be expressed in two values: (a) hydraulic EBCT at design flow and (b) average EBCT which includes also the regular down time of the plant. This value is calculated from the annual supply and the bed volume. Figure 9 displays both EBCT values from 24 treatment plants. The hydraulic EBCT at design flow of the plants varies between 3 and 10 minutes, the mean value is 4.2 minutes. The average EBCT varies between 4 and 17 minutes, with a mean of 10 minutes. The small EBCT leads to quite small plant designs and footprints, which allow to reduce installation costs. The As concentrations are between 10–40 ␮g/L in the raw water. The raw water pH values are in the range of 6.5–8 for the drinking water applications. The treatment capacities, expressed as

Adsorption density, g/kg

100 q(As V) q(As III) 10

1

0,1 5,5

6,5

7,5 pH

8,5

9,5

Figure 6. Adsorption densities with granular ferric hydroxide for As (III) and As (V) in a typical groundwater at different pH values.

197

Figure 7. Scheme for an As removal plant with granular ferric hydroxide (GEH®) and As treatment plant with granular ferric hydroxide, design flow 45 m3/h, installation with two vessels in parallel operation.

Minutes 16,0 14,0

Hydraulic EBCT Average EBCT

12,0 10,0 8,0 6,0 4,0 2,0 DEH VIL KSO TRH SUH BOW OSW LIN MIS ELK BGA WAK SGL MAH MAA NOB WHA SK-B RGL M3 M4 M9 MFH EGM

0,0

Figure 8. Empty bed contact time at design flow (hydraulic EBCT) and the yearly average value (average EBCT) (After Driehaus2002, updated).

bed volumes (BV) treated until exhaustion vary between 50,000 BV and more than 200,000 BV. The term exhaustion, as used in this paper, is not a unique parameter for all plants but depends on the preferences of the clients. Some prefer to exchange the media 5 ␮g/L As, others operate the plants close to the standard of 10 ␮g/L. Figure 9 shows the treatment capacities plotted against raw As and against raw water pH. Lower pH-values correlate well with high treatment capacities, but there is obviously only a small tendency of decreasing capacities at high As concentrations. 198

300000

Treatment Capacity in BV

Treatment Capacity in BV

300000 250000 200000 150000 100000 50000

250000 200000 150000 100000 50000

0 0

10

20

30

40

50

0

6

6,5

Raw arsenic, µg/L

7 7,5 Raw water pH

8

Figure 9. Treatment capacities versus raw As and versus raw water pH of 24 treatment plants with granular ferric hydroxide. 20 Raw Water Treated Water Calculated

Arsenic, µg/L

15

10

5

0

0

50.000 100.000 150.000 200.000 250.000 Bed Volumes

Figure 10.

Treatment data from the full scale plant OSW at a raw water pH of 7.1 and calculated exhaustion.

4.2.1 Performance of arsenic treatment with granular ferric hydroxide In order to predict treatment capacities and performance of the applications with granular ferric hydroxide, we developed a mathematical tool to calculate this data from the raw water quality. It was the aim of this tool to use only a small set of quality parameters, which were always and easily available. From the view of practicability, including more parameters does not necessarily lead to more accurate results, because parameters are not always reliable in terms of analytical and sampling errors. For example, there is no detailed knowledge how dissolved organic carbon (DOC) affects As adsorption. On the other hand, we have reliable data about the important competitors in As adsorption. From this competing ions, phosphate as the most important competitor is included in the tool, whereas the influence of sulphate and fluoride are far behind and are not included. The influence of silica as a competing and interfering substance is not yet fully understood, as either adsorption or precipitation reactions are involved. Results from the kinetic model and monitoring data from treatment plants are presented in Figures 10 and 11. Figure 10 shows the data from plant OSW. Plant OSW has a raw water pH of 7.1, As 14 ␮g/L, phosphate 280 ␮g/L and a total hardness of 90 mg/L CaCO3. The hydraulic EBCT is 2.8 minutes, and the plant is in operation for 10 h per day. The initial breakthrough occurs at 50,000 BV and is somewhat earlier than predicted by the kinetic model, but the further increase is less steep. After 140,000 BV, the effluent concentration remains constant at 5 ␮g/L, which corresponds to 60% removal. This is in sharp contrast to the mathematical prediction, which 199

Arsenic, µg/L

20 Average Raw Water Treated Run 1 Treated Run 2 Calculated

15

10

5

0

0

50.000

100.000

150.000

Bed Volumes

Figure 11.

Treatment data from the full scale plant KSO at a raw water pH of 7.9 and calculated exhaustion.

shows the steepest increase at around 60 to 50% removal. Thus the treatment capacity is much higher than predicted. One lifetime cycle with the granular ferric hydroxide lasts between 2.5 and 3 years, as the municipality decided to keep the As in potable water below 6 ␮g/L. This behavior of a steady state adsorption can not be explained by adsorption alone, which is here understood as monomolecular covering of the internal surface. If an adsorbent has a fixed number of adsorption sites, the ongoing operation and removal consumes free adsorption sites and the number of remaining sites decreases. Thus, the further uptake should decrease and lead to increasing effluent concentrations. We conclude, that there is another mechanism behind the adsorption in a monomolecular layer on the adsorption sites. We assume either multilayer adsorption, surface precipitation or an uptake of adsorbed species into the hydroxide structure and the forming of a new mineral phase. Treatment results from two lifetime cycles in Plant KSO are displayed in Figure 11, together with a predicted exhaustion curve. Plant KSO has a raw pH of 7.9, average raw As of 18 ␮g/L and a raw phosphate of 200 ␮g/L. The plant has an EBCT at design flow of 7.4 minutes and is operated for 14 h per day. Initial effluent data are not completely reported from this plant, but the initial breakthrough is at 40,000 BV. This is followed by an increase up to 10–11 ␮g/L at 90,000 BV. Afterwards, the effluent concentration remains constant at this level, showing a steady state adsorption with 40% removal. The mathematical model predicts a breakthrough for 10 ␮g/L at 80,000 BV and a further increase of the effluent As. Although the operational and raw water data are quite different from plant OSW, we have the similar effect at plant KSO at a higher concentration. Future research on this steady state adsorption is needed, to understand this effect. Up to date, this effect has been seen also at other localities using granular ferric hydroxide and in some lab scale column tests. This examples show, that a kinetic model with a small number of input parameters is able to describe the exhaustion in full scale applications, but with a tendency to underestimation. 5

WASTE STREAM CONSIDERATIONS

The element As can not disappear during treatment and will be accumulated in sludges, adsorbents, ion exchangers or membrane concentrates. All treatment techniques must produce waste streams, that are enriched in As. Table 1 gives an overview of the waste producing characteristics and the usual way of handling and disposing. Residuals from the conventional coagulation with ferric salts, thickened sludge, and exhausted iron oxide based adsorbents contain As which is bound to the iron hydroxide surface. Iron hydroxide is stable in oxidic environments, but will be dissolved and decomposed in reducing environments, and the toxic ingredients will be dissolved. Following also the pH-dependence of As uptake of iron 200

Table 1.

Waste streams of As-removal techniques, their properties and ways of discharge and disposal. Toxic waste (passes TCLP test)* Treatment

As-removal technique

Type of waste stream

Coagulation/ filtration

Ferric sludge

Fluids Unknown Water content up to 97% Sensitive for redox changes

Activated alumina with regeneration

Alkaline and acidic fluids

Fluids

Iron oxide based adsorbents

Exhausted adsorbent

Conventional disposal

Alternative disposal

Dewatering, drying

Sewer, landfill after dewatering

Brick manufacturing**)

Neutralization, precipitation with ferric salts

Treated liquids: Sewer Residual: landfill?

Solids, Passes water ⬍50% TCLP test Sensitive for redox changes

None

Landfill

Brick manufacturing*) Immobilization and utilization

Fluids

Not applicable

Precipitation with ferric salts

Treated liquids: Sewer or brine discharge line Residuals: landfill?

Efforts for brine recycling to minimize amounts

Concentrate Fluids liquids

Not applicable

Not used

Sewer or brine discharge?

Ion exchange Liquid, high saline brines

Membrane techniques: RO, NF

Properties

Not applicable

* TCLP . Toxicity Characteristic Leaching Procedure (US-EPA). ** Rouf & Hossain 2003.

oxide based adsorbents, As will be desorbed, if the pH of the surrounding solution is increased. Förstner & Haase (1998) concluded, that the consequence for a safe disposal is to keep this materials under oxidizing and neutral conditions. A disposal together with municipal waste, rich in organics will undoubtedly produce As-rich leachates, as well will do the co-disposal with caustic materials like fly ash. There is no common rule how to handle with the As treatment wastes. The conventional and alternative disposals given above are only a suggestion for more or less safe disposal or discharge. As it seems that the decision for one As treatment technique will be quite difficult, it will be as difficult to define the way of a safe and environmentally friendly waste disposal. In case of failing, one may create a bigger problem anywhere, either downstream in a supply chain or “downstream” in the environment. 6

CONCLUSION

The introduction into the aquatic chemistry of As showed that oxidation techniques for As(III) have to be evaluated in each individual case. Most of the presented, conventional and new treatment techniques have a better performance with As(V) in the raw water. Several proven oxidation methods exist, either by addition of chemicals or as fixed bed technologies. Application of strong oxidants like chlorine and ozone have side effects like formation of oxidation byproducts. One has to decide whether this is acceptable for a given application. Chlorine and hypochlorite seem to be the most frequently used oxidants for As(III) oxidation, because most waterworks staff are used to 201

handle chlorine. On the other hand, some countries (i.e. Germany) refuse to use chlorine for oxidation purposes. The decision on a suitable treatment method for As removal has to consider not only the raw water profile, treatment goals and flow rate, but also the infrastructure, availability of trained staff and raw materials, as well as disposal of waste streams. Usually this concentrates the choice to 2 or 3 different removal technologies. Coagulation followed by filtration is one of the conventional treatment techniques, which is especially suitable for large drinking water plants with well trained staff. Coagulation is based on dosing ferric or ferrous salts as coagulant, which precipitates, adsorbs As and builds up flocs, which are removed by filtration. Coagulation is in most cases effective to remove As, and typical doses of iron are in the range of 1–10 mg/L. As(III) is less efficiently removed and requires oxidation upstream. This technique requires an additional treatment for the As loaded backwash sludge, at least a dewatering. Coagulation could achieve low operation costs, as the ferric solutions are usually quite cheap. It is rarely applicable for small remote facilities, because the handling and dosing of corrosive chemicals requires a minimum infrastructure and trained staff. Nevertheless, there is a “manual” variation of coagulation applied in Bangladesh. This technique is applied by batchwise adding the chemicals packed for such a small volume to raw water in a bucket, stirring and waiting, and then filtering the water through sand in a second bucket. Ion exchange is given some attention in the USA, but bed life cycles of about 300–1000 BV between regenerations cause a lot of costly maintenance work for regeneration. Ion exchange as well as membrane techniques remain a domain of small household units, which could be operated automatically. Both are not selective to As and remove other anions or, with reverse osmosis, lead to a demineralised water. Activated alumina has been applied for As removal since more than 20 years. It exhibits bed life cycles of 1000–10000 BV until it needs to be regenerated or disposed. As activated alumina is sometimes available from local manufacturers at reasonable cost, it could be the method of choice for small facilities, especially those with neutral to acidic raw waters and As(V). The new techniques with iron oxide based adsorbents, which were developed and introduced into the market during the last 8 years could overcome the shorts of the previously mentioned techniques. The actually best medias have bed life cycles of 100,000 to 300,000 BV until exhaustion. Exhausted adsorbents are disposed of and replaced by factory fresh adsorbent. After 7 years of experience with granular ferric hydroxide they have proven their high performance under various conditions in Europe and overseas. The operation and maintenance of treatment plants is quite simple and fulfills all expectations of a simple adsorption process. No chemicals need to be dosed to the water and the only task for maintenance is backwashing of the adsorber vessel on a monthly frequency. All drinking water standards were met in all plants and the breakthrough of As in the treated water is very slow. The operational cost for treatment plants with iron based adsorbents are mostly comparable to coagulation plants, and for small facilites they could be the most economic solution. An interesting variation of iron based adsorbents is the use of natural and locally occurring iron oxides and hydroxides. Singh et al. (1988) investigated As(III) removal by hematite ores. Another variation is the use of metallic iron, as proposed by Khan et al. (2000). Metallic or zerovalent iron oxidizes in the raw water and the developed iron hydroxides acts as an adsorbent for As. During the last years, especially driven by the implementation of a new As standard in the USA, came up a number of new adsorption medias for As removal. Some are still under development or only lab tested, others are commercially available. Also some of the offered adsorbent medias are not really suitable for As removal, others have been proven under pilot plant conditions to have a suitable performance. Iron oxide based adsorbents will probably play an important role in As removal in the future.

REFERENCES Anderson, L.D.C. & Bruland, K.W. 1991. Biogeochemistry of arsenic in natural waters: The improtance of methylated species. Environ. Sci. Technol. 25(3): 420–427.

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Baldauf, G. 1995. Aufbereitung arsenhaltiger Grundwässer. GWF Wasser Special 136(14): 99–110. Borho, M. 1996. Arsenentfernung in Grundwasserwerken durch optimierte Kopplung von Oxidations und Fällungs-/Flockungsverfahren. Bericht aus der Wassergüte- nd Abfallwirtschaft der Technischen Universität München, 127 (Doctoral thesis). Chang, J. 2004. Demonstration of emerging technologies for arsenic removal. Research Foundation Report. 119p. AWWARF-AWWA, Denver. Driehaus, W., Seith, R. & Jekel, M. 1995. Oxidation of arsenate(III) with manganese oxides in water treatment. Water Research 29(1): 297–305. Driehaus, W., Jekel, M. & Hildebrandt, U. 1998. Granular ferric hydroxide–a new adsorbent for the removal of arsenic from natural water. Jour. Water SRT-Aqua 47: 30–35. Ferguson, J.F. & Gavis, J. 1972. A review of the arsenic cycle in natural waters. Water Res. 6: 1259–1274. Förstner, U. & Haase, I. 1998. Geochemical demobilization of pollutants in solid waste–implications for arsenic in waterworks sludges. Journal Geochemical Exploration 62: 29–36. Frank, P. & Clifford, D. 1986. Arsenic(III)-oxidation and removal from drinking water. USEPA 600-52-86/021. Jekel, M.R. 1994. Arsenic removal in water treatment. In: Nriagu J. (ed.) Arsenic in the Environment, Part 1, Cycling and Characterization: pp 433–446, John Wiley, New York. Jekel, M. & Seith, R. 1999. Comparison of conventional and new techniques for the removal of arsenic in a full scale water treatment plant.–World Water Congress 1999, Buenos Aires, Proceedings. Khan, A.H., Rasul, S.B., Munir, A.K.M., Habibuddowla, M., Alauddin, M., Newaz, S.S., Hussam, A. 2000. Appraisal of a simple arsenic removal method for groundwater of Bangladesh. J. Environ. Sci. Health, Part A: Toxic/Hazardous Substances and Environmental Engineering 35(7): 1021–1041. Kim, J., Benjamin, M.M., Kwan, P. & Chang, Y. 2003. A novel ion exchange process for As removal. Journal American Water Works Association 95(3): 77–85. Pierce, M.L. 1981. Chemical modeling of arsenic in aqueous systems: Ph.D. Thesis, Arizona State University. Rouf, M.A. & Hossain, M.D. 2003.Effects of using arsenic iron sludge in brick making. Presented at the BUET-UNU International Symposium Fate of Arsenic in the Environment, February 2003 Dhaka, Bangladesh. URL: http://www.unu.edu/env/Arsenic/BUETSymposiumProc.htm. Seith, R. & Jekel, M. 1997. iologische Oidation von Arsen(III) in Festbettreaktoren. Vom Wasser 89: 283–296. USEPA 2000.Technologies and costs for removal of arsenic from drinking water, EPA 815-R-00-028. Singh, D.B., Prasad, G., Rupainwar, D.C. & Singh, V.N. 1988. As(III)-removal from aqueous solution by adsorption. Water Air Soil Pollution 42: 373–386. Waltham, C.A. & Eick, M.J. 2002. Kinetics of arsenic adsorption on goethite in the presence of sorbed silicic acid. Soil Sci Am J. 66: 818–825. Waychunas, G.A., Rea, B.A., Fuller, C.C. & Davies, J.A. 1993. Surface chemistry of ferrihydrite: Part 1. EXAFS studies of the geometry of coprecipitated and adsorbed arsenate. Geochim. Cosmochim. Acta 57: 2251–2269. WHO 2001. United Nations synthesis report on arsenic in drinking water. http://www.who.int. Zeng, L. 2002. Preparation of an iron(III) oxide adsorbent for arsenic removal. URL: http://www. globe2004.com/2002/presentations/zeng.doc (Accessed on October 4, 2004).

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Natural enrichment of arsenic in a minerotrophic peatland (Gola di Lago, Canton Ticino, Switzerland), and implications for the treatment of contaminated waters Z.I. González-A., M. Krachler, A.K. Cheburkin & W. Shotyk Institute of Environmental Geochemistry, Heidelberg University, Heidelberg, Germany

ABSTRACT: Minerotrophic peatlands are natural sinks for arsenic (As) in groundwaters: they remove As from input waters as they accumulate As in peat. In Gola di Lago (GdL), Canton Ticino, Switzerland, As concentrations in peat are highly elevated (ca. 150x) compared to “background” values. Stream waters entering this peatland contain up to 400 ␮g/L As, but the waters leaving the peatland contain less than 2 ␮g/L. The rates and mechanisms of this process have not yet been studied, but offer great promise for the low-cost treatment and remediation of As-bearing waters by natural wetlands. The main goal is to understand how As is removed from the incoming fluid by the peat. We present here a panorama of our research, methodology, previous results and future work.

1

INTRODUCTION

Elevated concentrations of arsenic (As) in groundwaters represent a potential health hazard in many countries, including: Argentina, Bangladesh, China, Ghana, Hungary, India, Laos, Mexico, Mongolia, Myanmar, Nepal, Pakistan, Sumatra, Taiwan and Vietnam (Bhattacharya et al. 2002, Berg 2003). Arsenic is a metalloid widely distributed in the earth’s crust with a typical abundance on the order of 2 mg/kg. It occurs in trace quantities in all rocks, soils, waters and air. Arsenic can exist in four valences states: ⫺3, 0, ⫹3 and ⫹5. Under reducing conditions, arsenite (As III) is the dominant form; arsenate (As V) is generally the stable form in oxygenated environments. Elemental arsenic is not soluble in water. Arsenic salts exhibit a wide range of solubilities depending on pH and the ionic strength of the solution (Gomez-Caminero et al. 2001). Arsenic is commonly enriched in peat and coal (Valkovic 1983, Swaine 1990). Peat is fossil plant matter, which forms and accumulates in water – saturated, anoxic wetlands. In the absence of oxygen, bacteria use Mn, Fe, and S as terminal electron acceptors, in addition to a variety of trace elements. The strongly reducing condition of these waters, combined with the abundance of organic matter capable of complexing trace metals, creates a natural geochemical trap for a broad range of trace elements. Peatlands containing up to 10 weight percent Cu and up to 3% U have been reported (Shotyk 1988) but As, Hg and Se also may accumulate in peat (Shotyk et al. 1992, Shotyk et al. 2003). Gola di Lago is a small minerotrophic (i.e. water–fed) peatland in Canton Ticino, Switzerland. A core taken from this peatland showed pronounced enrichments of As, Se and U. This peat core also showed a strong enrichment of S and depletions of Fe and Mn (Shotyk 1996, unpubl.). In southern Switzerland, several areas have been found where the concentrations of arsenic in rocks, soils and waters are higher than recommended guideline values (Pfeifer et al. 2002). The source of this contamination seems to be natural and attributable to rock weathering. 205

The environmental chemistry of arsenic is complex and varies from region to region, depending on unique, local characteristics. There are many investigations about arsenic and its behavior in different environments and under different conditions, however, its behavior in peat has not been studied in detail. We are in the process of beginning a study of the rates and mechanisms of As accumulation in a peatland (Gola di Lago) in Canton Ticino, Switzerland. Our main goal is to quantify the rates and understand the mechanisms of arsenic enrichment in this peatland. In this paper we present some preliminary results, methodology and future work.

2

MATERIAL AND METHODS

The study includes careful examination of both the aqueous phase (waters flowing into and out from the peatland) and the solid phase (peat, sediments and minerals found therein). 2.1

Aqueous phase

The water which interacts with the peatland was analyzed mainly for total dissolved As using SFICP-MS (Krachler et al. 2002). The samples were taken from 5 different locations, one from each stream and two more from the peatland (Fig. 1). 2.2

Solid phase

One peat core was taken in 2001 from the Gola di Lago peatland, Canton Ticino, Switzerland, as the map shows (Fig. 1). This peat core was frozen before cutting into 1 cm slices: the edges of each slice were trimmed and the remainder was cut in half. One half was archived and the other half was divided into eight similar parts of known volume; two of these parts were removed for later Hg analyses and the rest were weighed, dried and milled to a fine powder using an agate ball mill. The samples were analyzed for As, Fe, Mn, S, Se, U, and Ti using non-destructive XRF analyses (Cherburkin & Shotyk 1996); this method allows a large number of samples to be measured in the shortest possible time, at low cost. Using this approach, As concentrations are measured as the difference between (Pb ⫹ As) and (Pb). One consequence of this is that the accuracy of As data is poor in samples containing high Pb concentrations, and low As. However, the method does provide a rapid illustration of the distribution of As and its enrichment in the peat, relative to crustal rocks.

Stream A

Stream B

Peatland water sampling 1

Peat core 2001 Peatland water sampling 2

Peatland Stream C

Figure 1.

Map of the peatland in Gola di Lago, Switzerland (main site of interest).

206

3 3.1

RESULTS AND DISCUSSION Aqueous phase

The results of the water analyses results are listed in Table 1. Concentrations of Fe, Mn and U determined in the water samples by SF-ICP-MS were at least 500 times (As ⬇ 100 times) higher than the corresponding detection limits. To check the accuracy of the results the riverine water reference material SLRS – 4 (Natural Research Council of Canada), has been analyzed for quality control purposes. Streams A and B both enter the peatland and C is the outlet. Peatland water 1 and 2 are water samples taken at the surface of the peatland. The results show that the concentration of arsenic changes drastically, from the inlet (⬇400 ␮g/L) to the outlet (⬇2 ␮g/L). Stream B is enriched with uranium too. Fe and Mn will be studied in more detail later, using “peepers” (Steinman & Shotyk 1996).

3.2

SOLID PHASE

It appears that sulphate ([SO4]2) and arsenate ([AsO4]) in the incoming waters are reduced to sulphide ([HS]-) and arsenite ([H3AsO3]0) in the pore waters of the peatland; these reactions, combined with the reductive dissolution of Fe oxides present in the peat, might allow Fe, S, and As concentrations to reach saturation with respect to arsenopyrite [FeAsS](s). Other possible reduction products are As (elemental), AsS (realgar) and As2S3 (orpiment). The main reactions can be summarized as:

And the reduction products reactions as:

Table 1.

Results (␮g/L) of the water samples with the SF-ICP-MS.

Sampling place

As

Fe

Mn

U

Stream A Stream B Stream C Peatland – water 1 Peatland – water 2

0.50 ⫾ 0.02 408 ⫾ 34 1.8 ⫾ 0.13 0.55 ⫾ 0.027 0.37 ⫾ 0.026

11.9 ⫾ 0.4 9.14 ⫾ 2.0 10 ⫾ 0.30 12.7 ⫾ 0.386 5.09 ⫾ 0.16

0.10 ⫾ 0.02 0.20 ⫾ 0.05 0.09 ⫾ 0.006 0.18 ⫾ 0.009 0.04 ⫾ 0.004

0.02 ⫾ 0.0001 1.6 ⫾ 0.054 0.04 ⫾ 0.001 0.03 ⫾ 0.0001 0.02 ⫾ 0.0001

207

0

-20

-20

-40

-40

Depth (cm)

Depth (cm)

0

-60

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-100

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10

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30

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120 140

160 180

60

S (%)

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Depth (cm)

As (ppm)

-60

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-100

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-120 0

20

40

60

80

100 120 140 160 180 200

Fe (%)

Figure 2.

0

200

400

600

800

1000

Mn (ppm)

Peat core profiles showing the distribution of As, S, Fe and Mn determined with XRF (EMMA).

The results shown in Figure 2 indicate that only a small portion of the As be corresponds to the enrichments of Fe and Mn at a depth of 20 cm. The hypothesized formation and accumulation of arsenopyrite in peatlands represents a natural geochemical filter, which removes As from water. Once this process is better understood, it will be helpful in applying the extensive natural wetland areas for treating and remediating As-rich waters. The results shown here indicate that the peatland also traps U. Wetlands such as the peatland described here, therefore could also be helpful in treating U bearing wastewaters. The profiles for Se and U are shown in Figure 3. The enrichments of Se and U suggest that the peat is also reducing with respect to Se(VI) and U(VI). With respect to selenium, it is often present in uranium deposits (Strawn et al. 2002). The main reactions can be summarized as:

The profiles for Ti, and the profiles of As/Ti and Se/Ti ratios are shown in Figure 4. Titanium is found in insoluble minerals in the environment (Cox 1995) and its abundance in the peat reflects the concentration of mineral material present in peatlands (Shotyk 1988). 208

0

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Depth (cm)

Depth (cm)

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1

2

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4

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0.5

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Se (ppm)

Depth (cm)

Figure 3.

2.0

2.5

U (ppm)

Peat core profiles showing the distribution of Se and U determined with XRF (EMMA).

0

0

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200 400 600 800 100 1200 1400

Ti (ppm)

-120 0

50

100 150 200

Se EF

250 300

350

0

50

100

150

200

250

As EF

Figure 4. Peat core profile for Ti and graphics for Se/Ti and As/Ti ratios as an estimate of the extent of As and Se enrichment.

To estimate the extent of As and Se enrichment, the As/Ti and Se/Ti ratios of the peat samples were normalized to their corresponding ratios in crustal rocks (As 2 ppm, Se 0.083 ppm and Ti 3117 ppm. (Wedepohl 1995). The enrichment factor (EF) profiles show that high As and Se concentrations in the peatland are not a simply abundance of mineral particles. In fact, the peat samples are enriched in As (31 to 200x) and Se (37 to 340x), far out of proportion with the amount of mineral matter.

4

CONCLUSIONS

The minerotrophic peatland from Gola di Lago, Canton Ticino, Switzerland is removing arsenic, selenium, and uranium from stream water, with respect to As. Flow rate data and more peat core data are needed to obtain a mass balance. The concentrations of As and Se will be quantified using HG-AFS to obtain an accurate assessment of the removal process. Our main challenge is to understand the mechanism of arsenic accumulation in peatland. The results obtained to data suggest that Se and U are similarly affected. 209

REFERENCES Berg, M. 2003. Arsenic in drinking water, Vietnam a new focus of attention. EAW AG News 53:12–15. Bhattacharya, P., Frisbie, S.H., Smith, E., Naidu, R., Jacks, G. & Sarkar, B. 2002. Arsenic in the Environment: A Global Perspective. In B.Sarkar (ed.) Handbook of Heavy Metals in the Environment (Chapter 6), New York: Marcell Dekker Inc., pp. 145-215.Cheburkin, A.K. & Shotyk, W. 1996. An Energy-dispersive miniprobe multielement analyzer (EMMA) for direct analysis of Pb and other trace elements in peats. Fresenius Journal of Analytical Chemistry 354: 688–691. Cox, P.A. 1995. The elements on earth, inorganic chemistry in the environment. Oxford University Press Inc. New York. Gomez, A.C., Howe, P., Hughes, M., Kenyon, E., Lewis, D.R., Moore, M., Ng, J., Aitio, A. & Becking, G. 2001, Health Criteria 224: World Health Organization. Krachler, M., Mohl, C., Emos, H. & Shotyk, W. 2002. Analytical procedures for the determination of selected trace elements in peat and plant samples by inductively coupled plasma mass spectrometry. Spectrochimica Acta B 57: 1277–1289. Pfeifer, H.R., Beatrizotti, G., Berthoud, J., De Rossa, M., Girardet, A., Jäggli, M., Lavanchy, J.C, Reymond, D., Righetti, G., Schlegel, C., Schmit, V. & Temgoua E.. (2002). Natural arsenic-contamination of surface and ground waters in southern Switzerland (Ticino). Bull. Appl. Geol. 7(1): 81–103. Shotyk, W. 1988. Review of the organic geochemistry of peats and peatland waters. Earth Science Reviews 25: 95–176. Shotyk, W., Nesbitt, H.W. & Fyfe W.S. 1992. Natural and antropogenic enrichments of trace metals in peat profiles. International Journal of Coal Geology 20: 49–84. Shotyk, W., Goodsite, M.E., Roos-Barraclough, F., Frei, R., Heinemeier, J., Asmund, G., Lohse, C. & Hansen, T.S. 2003. Anthropogenic contributions to atmospheric Hg, Pb and As deposition recorded by peat cores from Greenland and Denmark date using the 14C ams “bomb pulse curve”. Geochimica et Cosmochimica Acta 67(21): 3991–4011. Steinman, P. & Shotyk, W. 1996. Sampling of in situ filtered pore waters in peat lands using “peepers”. Fresenius Journal of Analytical Chemistry 354: 709–713. Strawn, D., Doner, H., Zavarin, M. & McHugo, S. 2002. Microscale investigation into the geochemistry of arsenic, selenium and iron in soil developed in pyretic shale materials. Geoderma 108: 237–257. Swaine, D.J. 1990. Trace elements in coal. Butterworth, London. Valkovic, V. 1983. Trace elements in coal. CRC press, Boca Raton (2 Vols.). Wedepohl, K.H. 1995. The composition of the continental crust. Geochimica et Cosmochimica Acta 59(7): 1217–1232.

210

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

A comparative study for the removal of As(III) and As(V) by activated alumina Tony Sarvinder Singh & K.K. Pant Department of Chemical Engineering, Indian Institute of Technology, New Delhi, India

ABSTRACT: Contamination of arsenic in drinking water is a severe health risk in different parts of the world. In the present study effectiveness of activated alumina (AA) was examined for the adsorption of arsenite and arsenate ions from water. Batch experiments with temperature and agitation control was employed to determine the adsorption capacity and adsorption kinetics for arsenite and arsenate removal. Removal of both As(III) and As(V) was found to be strongly dependent on initial pH of water. Effects of contact time, temperature, initial concentration of adsorbent dose and adsorbate concentration were studied to establish optimum conditions. The present investigation revealed that arsenate ions adsorb well on activated alumina at pH less then pHZPC (8.2) and maximum removal (98.6%) was achieved at pH 5.8 from water. Compared to As(V), As(III) ions have high affinity towards activated alumina at pH 7.6. The experimental data for adsorption of both As(III) and As(V) fitted well to the Freundlich and Langmuir adsorption isotherm and adsorption was favorable. Column experiments revealed that the breakthrough performance strongly depends on bed height and feed flow rate. By taking chemistry of water into consideration, this process can be used effectively for the removal of arsenic species.

1

INTRODUCTION

Inorganic pollutants in particular heavy metal ions constitute a major class of water contaminants Arsenic is the one of the abundant element in the earth’s crust (Cullen & Reimer 1989). Arsenite [As(III)] and arsenate [As(V)] are highly toxic inorganic species that represent a potential threat to the environment and human health. Arsenic in small quantities is necessary as a nutrient to humans but if ingested in too large dose (⬎LD50) or ingested in smaller quantities over a long period of time, it can cause various carcinogenic effects. Arsenic chemistry in aquatic environment is quite complicated because the metal can be stable in four oxidation states ⫹5, ⫹3, 0 and ⫺3 under different redox conditions. At the high Eh value encountered in oxygenated waters, arsenic 2⫺ 3⫺ acid species (H3AsO4, H2AsO⫺ 4 , HAsO4 and AsO 4 ) are stable whereas under mild reducing conditions arsenous acid species (H3AsO3, H2AsO⫺3 and HAsO32⫺) become stable (Ferguson & Gavis 1972). In the presence of sulfide, realgar (AsS) and orpiment (As2S3) may be present as stable solids at pH less than about 5.5 and at Eh value of about 0 volts. HAsS2 and AsS are the predominant aqueous species. Under extremely reducing environment, arsenic may occur in traces being only slightly soluble (Ghosh & Yuan 1987). Toxicity of the arsenic varies greatly according to its oxidation state as As(III) is much more toxic than As(V). Acute and chronic arsenic poisoning in drinking water has been reported worldwide (Dutta et al. 1997). Although environmental regulations have limited the production and uses of arsenic and its compounds, they are extensively used in metallurgy, agriculture, forestry, electronics, pharmaceuticals and glass and ceramic industries. It is introduced into the environment through weathering of rocks and mine tailings, industrial wastes discharge, fertilizers, agricultural employments of pesticides, smelting of metals and burnings of fossil fuels. The presence of arsenic in drinking 211

water has caused acute and chronic poisoning. Several diseases have been linked to the consumption of arsenic containing ground water in Bangladesh, Taiwan, Argentina, Mexico, Chile, China, Hungary, Thailand, USA, New Zealand, South Africa and in India. Arsenic has been classified as a carcinogen by the USEPA (USEPA 2001). Its presence in drinking water increases the risk of skin, lung, kidney and bladder cancer. Other than cancer many diseases like haff’s disease, Blackfoot disease etc. may result from arsenic consumption (Ng et al. 2003). It has been reported that longterm uptake of arsenic contaminated drinking water has produced gastrointestinal, skin, liver and nerve tissue injuries. If arsenic builds up to higher toxic level, open lesions, organ damages, neural disorder can result (Morales et al. 2000). Taking into consideration of its health effects and toxicology, occurrence and human exposure, some regulatory agencies have revised the maximum contaminant level for arsenic in drinking water from 0.05 mg/L to 0.01 mg/L. In 1993, WHO recommended a guideline of 0.01 mg/L. Various physical and chemical techniques have been applied for the removal of arsenic including complexation with polyvalent metal species salts as ferric ion, co precipitation with amorphous hydrous metal oxides (Gulledge & O’Conor 1973, Pierce & Moore 1980), lime softening (Lee & Rosehart 1972), ion exchange and activated alumina (Rosenblum & Clifford 1984, Bellack 1971). Most of these studies are limited for the removal of As(V) species. The conventional methods of arsenic removal such as coagulation precipitation with lime, alum and ferric sulfate produces a wet bulky sludge and disposal of this is a matter of concern as it contains very high arsenic concentration. Adsorption systems are rapidly gaining prominence as treatment processes which produce good quality water. Adsorption has also been found effective for arsenic removal. Various adsorbents for arsenic removal have been attempted which include amorphous iron hydroxide (Pierce & Moore 1982), activated carbon (Gupta & Chen 1978, Huang & Fu 1984, Pattanayak et al. 2000), activated alumina (Bellack 1971, Ghosh & Yuan 1987), hydrous zirconium oxide (Suzuki et al. 1997), hematite and feldspar (Singh et al. 1996), industrial waste (Low & Lee, 1995), lanthanum-loaded silica gel (Wasay et al. 1996), metal-loaded coral limestone (Ohki et al. 1996), coconut husk (Manju et al. 1998) and biological materials such as living or non-living biomass. Chitin, chitosan (Elson et al. 1980, Muzzarelli et al. 1984). However many of these adsorbents are suitable for As(V) removal and also require technical expertise for their impregnation. Activated alumina has been found effective for removal of various organic and inorganic impurities from water and wastewater. The objective of this work is to develop a methodology concerning the adsorption As(III) and As(V) onto activated alumina (herein after referred to as AA) using batch and continuous operation have been discussed taking into consideration the variables such as pH, feed flow rate and contact time.

2

MATERIALS AND METHOD

All chemicals used were of analytical grade. Arsenite and arsenate stock solutions were prepared from analytical grade dehydrated sodium arsenite (NaAsO2) and sodium arsenate (Na2HAs2O7⭈7H2O, Merck, Germany) respectively. Dissolution of arsenite and arsenate salts also includes addition of HCl. Further working solutions of different concentration were prepared by diluting the stock arsenic solution in de-ionised water. Commercially available granular AA (IPCL Baroda, India) was used in this study. AA was pretreated by oven drying at 378 K for 24 h, and then stored in a desiccator for further analysis and experiments. Surface area and pore volume of the adsorbent was measured using ASAP 2010 micro pore surface area analyzer (Micromeritics Corporation, USA). Nitrogen gas at liquid N2 temperature (77 K) was used as analysis gas. Prior to the multi point BET analysis, the sample was degassed in a sample tube at 473 K for 10 hours to remove volatile impurities and moisture, which may have adsorbed on the surface previously. The surface area and pore volume of the activated alumina was 212

Table 1.

Properties of activated alumina.

Properties

Quantitative value

Surface area (m2/gm) Particle size (spherical mm) Pore volume (cc/gm) Bulk density (kg/m3) Loss on attrition (%) Al2O3 (%) Na2O (%) Fe2O3 (%) SiO2 (%)

370 2.0 ⫾ 0.1 0.42 800 0.2 93.1 0.02 0.04 0.03

found to be 370 ⫾ 2 m2/g and 0.42 ⫾ 0.03 cm2/g respectively. The properties of activated alumina are listed in Table 1. Performance runs were carried out at various pH in the range of 2–13 to investigate the effect of initial solution pH on removal of As(III) and As(V). Further studies were conducted at the pH where maximum As(III) and As(V) removal was observed. Equilibrium data were collected by taking 100 ml solution of As(III) and As(V) of known concentration (0.5 and 1.5 mg/L) in a series of reagent flasks, which were placed on a shaker at 85 rpm for 24 hours. To each flask, different AA doses (0.1–1.3 g) were added. After 24 hours, samples were filtrated through Whatman filer paper no 1 and analyzed for residual As(III) and As(V) concentration. Batch kinetic studies were conducted at two different initial As(III) and As(V) ion concentration. 1.0 g of AA was suspended in 100 ml of As(III) and As(V) solution of known initial sorbate concentration and the pH was adjusted with 0.01 N NaOH/HCl. The mixture was continuously stirred at 85 rpm. Samples were withdrawn at pre-determined time intervals in the range of 1–24 hours, filtered and analysed for residual As(III) and As(V) ion concentration. Column experiments were conducted at carious bed heights of 6, 12 and 18 cm corresponding to residence time of 4.7 minutes, 9.4 minutes and 18.8 minutes respectively. As(III) and As(V) ions were passed through the column at a flow rate for 2 ml/min with an initial metal loading of 0.5 mg/L. Samples were collected at various time intervals and analysed till the exit sorbate concentration equal to inlet arsenic concentration. Arsenic content was determined by Graphite tube–Atomic Absorption Spectrophotometer (Varian Spectra AA-860, Varian Australia) by using hollow cathode lamp at a wavelength of 193.7 nm. The instrument was calibrated by using As(III) and As(V) standard solutions of 25, 50 and 100 ␮g/L separately which were prepared in diluted HNO3. 3

RESULTS AND DISCUSSION

3.1 Effect of initial solution pH on As(III) and As(V) adsorption The adsorption of As(III) and As(V) on AA was found to be strongly dependent on the initial pH. Maximum As(III) and As(V) removal of 96.2% and 98.6% were observed at pH 7.6 and pH 5.8 respectively for an initial concentration of 0.5 mg/L. High sorption capacity of activated alumina can be attributed to its high surface area, zero point of charge and stronger interactions. Arsenate ions were removed better than arsenite ions (Fig. 1) indicating the better binding of these ions on AA surfaces. The pH of the solution determines the concentration distribution of the ionic forms of the As(III) and As(V). In pH range between 3 to 6, As(V) ion occurs mainly in the form of H2AsO4 while a divalent anion HAsO42⫺ dominates at higher pH values (8 to 10.5). In the intermediate region, i.e. in pH range 6 to 8, both species can co-exist. It is evident from the results that AA effectively adsorbs both species of H2AsO4 and HAsO42⫺. For As(V) removal, optimum pH range for different sorbents reported in literature were 6–8 (La(III) and Y(III) impregnated alumina), 213

100 % Removal

80 60 40 As(V)

20

As(III)

0 0

2

4

6

8

10

12

14

pH

Figure 1.

Effect of pH on As(III) and As(V) removal (C0 0.5 mg/L; As dose 10 g/L; T 298 K].

2–4 (molybdate impregnated chitosan beads) (Wasay et al. 1996, Tokunaga et al. 1997, Dambies et al. 2000). In the present investigation, pH value of 7–8 and 5–6 were found most effective for As(III) and As(V) respectively. Since point of zero charge (pHpzc) for different type of alumina is around 8.4–9.1 (Stumm & Morgan, 1981, Bowers & Huang 1985). Surface of AA is positively charged until the pH reaches the point of zero charge. The anionic species, arsenate, would thus have stronger interaction (specific binding) with AA surface and have higher uptake. In the present investigation uptake of arsenic species with AA first increases with increase in pH up to pH 7.6 and 5.8 for As(III) and As(V) respectively and subsequently dropped significantly at higher pH (Fig. 1). For pH less than 8.0, the surface of AA is predominantly positively charged and the major arsenic species is H2AsO4, and HAsO⫺ 3 . Therefore, specific binding is expected for the adsorption process. As pH increases, the portion of positively charged surface sites on AA decreases, causing the reduction of adsorption. The adsorption of As(III) and As(V) is governed by both the surface charge of AA and the form of arsenic species in the water. As a result of that, pH is a strong factor in the uptake of arsenite and arsenate by activated alumina. 3.2 Kinetics study Kinetics experiments show that the adsorption of both As(III) and As(V) increases with the lapse of time. The maximum uptake of As(III) and As(V) over AA was 230 mg/kg and 450 mg/kg respectively. The adsorption of As(III) was rapid during the first six hours as compared to four hours in case of As(V) after which the rate slowed down. The equilibrium approached almost at the same time (approx. 9–10 hours) for both As(III) and As(V) respectively. At higher concentrations of both As(III) and As(V) lower percentage removal was obtained. The results obtained from the experiments were used to study the rate-limiting step in the adsorption process. The adsorption rate constant (kad) for adsorption was determined from the following first order rate expression (Gupta et al. 1988) (1) where qe and qt (both in mg/g) are the amount of arsenic adsorbed per unit mass of activated alumina at equilibrium and at time t respectively and kad is the rate constant (l/hr). The value for kad was calculated from the slope of the linear plot of log (qe ⫺ q) versus time (Fig. 2). The ad sorption rate constants are reported in Table 2. Rate constants were found independent of initial As(III) and As(V) concentration in the range of experimental conditions studied indicating that the adsorption of As(III) and As(V) on activated alumina follows first order rate kinetics. In case of strict surface adsorption, a variation in rate should be proportional to the first power of concentration. However, when pore diffusion limits the adsorption process, the relationship between initial solute concentration and 214

0

As(V) 0.5 mg/l As(V) 1.5 mg/l As(III) 0.5 mg/l As(III) 1.5 mg/l

-1 ln(qe-q)

-2 -3 -4 -5 -6 0

Figure 2. Table 2.

2

4

6 8 Time (hr)

10

12

14

Lagergren plot for adsorption kinetics for As(III) and As(V) removal by AA. Rate kinetics and adsorption isotherm constants.

As(III) [C0 0.5 mg/L] [C0 1.5 mg/L] As(V) [C0 0.5 mg/L] [C0 1.5 mg/L]

Kad(hr1)

R2

Kp (mg g⫺1 hr⫺0.5)

0.217 0.231 0.235 0.253

0.90 0.97 0.985 0.981

0.02 0.11 0.03 0.12

Langmuir isotherm constants

N

K

b (mg/g)

qm (l/mg)

2.318

0.214

7.153

0.196

1.751

0.448

7.426

0.308

As(V) 0.5 mg/l As(V) 1.5 mg/l As(III) 0.5 mg/l As(III) 1.5 mg/l

0.2 0.16 qt (mg/g)

Freundlich isotherm constants

0.12 0.08 0.04 0 0.5

1.5

2.5

Time

Figure 3.

0.5

3.5

(hr0.5)

Intraparticle diffusion curves for As(III) and As(V).

the rate of adsorption will not be linear (Srivastva et al. 1997). Besides for adsorption on the outer surface of adsorbent, there is also a possibility of transport of adsorbent ions from the solution to the pores of the adsorbent due to stirring on batch process. This possibility was tested in terms of a graphical relationship between amount of As(III) and As(V) adsorbed and square root of time (Fig. 3). In order to show the existence of intraparticle diffusion in the adsorption process, the amount of As(III) and As(V) ions sorbed per unit mass of adsorbents at time t, qt was plotted as a function of square root of time, t0.5 (Fig. 3). The rate constant for the intraparticle diffusion was obtained using the equation (2) 215

where kp (mg/g hr0.5) is the intraparticle diffusion rate constant. The plot for intraparticle diffusion of As(III) and As(V) shows that initially curved portion reflects film or boundary layer diffusion effect and the subsequent linear portion attribute to the intraparticle diffusion effect (Knocke & Hemphill, 1981). The slower of the two processes will be rate-limiting step. Figure 3 depicts that intraparticle diffusion was slow and rate determining step for sorption of arsenic ions. kp values obtained from the slope of the linear portion of the curve were found to be 0.02 mg/g hr0.5 and 0.03 mg/g hr0.5 for As(III) and As(V) at initial metal concentration of 0.5 mg/L. These values were higher at higher arsenic concentration. This result suggests that the adsorption is governed by the diffusion within pores of the adsorbent. The linear portions of the curves do not pass through the origin (Fig. 3). This indicates that mechanism of both arsenite As(III) and As(V) removal on activated alumina is complex and both, the surface adsorption as well as intraparticle diffusion contribute to the rate determining step. 3.3 Equilibrium studies Equilibrium studies were conducted by varying AA doses (0.1–1.3 g/100 ml) for As(III) and As(V) (C0 1.0 mg/L) at 298 K. Percent arsenic removal was found to increase with the increase in adsorbent doses. Among several isotherm equations, which have been reported for modelling of adsorbtion systems, the Freundlich and Langmuir isotherm models were selected for the analysis of sorption data obtained at different sorbent concentration, keeping the sorbate concentration constant at 1.0 mg/L. The linearized plot of Langmuir (plot of 1/qe and 1/Ce) isotherm is given in Figure 4. The Langmuir isotherm can be described as follows (3) where Ce, x and m are sorbate concentration at equilibrium (mg/L), amount of As(III) or As(V) ions removed (mg/L) and mass of sorbent (g) respectively. The Langmuir constant qmax, defined as the amount of adsorbate per unit weight of adsorbent to form a complete monolayer on the sorbent surface was found to be 196 mg/kg and 308 mg/kg for As(III) and As(V) respectively, while, b which reflects quantitatively the affinity between the adsorbent and adsorbate was found 7.153 and 7.426 for As(III) and As(V) respectively. A higher value of b implies strong surface bonding and hence favorable adsorption. However compared to As(III), As(V) can be more easily removed from solution. Freundlich isotherm is mathematically represented as (4) where x/m is the amount of As(III) or As(V) ions adsorbed at equilibrium (mg/g), Ce the equilibrium concentration of As(III) or As(V) in solution (mg/L), K and 1/n the Freundlich constants are 16

1/qe (g/mg)

12 8 As(III)

4

As(V) 0 0

Figure 4.

5

10

15 20 1/Ce (l/mg)

25

30

Langmuir adsorption isotherm plot for As(III) and As(V) removal onto AA.

216

shown in Figure 5. The logarithmic plots of x/m versus Ce gave straight line which supports the applicability of the Freundlich isotherm model to the present study. The constants K and 1/n were evaluated from the intercept and the slope of the straight lines by least square method . The values are reported in Table 2. 3.4

Column experiments

Adsorption involves accumulation or concentration of an adsorbate species at an adsorbent surface, or interface. In a fixed bed adsorption system, the adsorbent located closest to raw water/feed inlet saturates first where maximum adsorption takes place initially. The adsorption zone progresses downwards as time passes and then approaches the exit of the fixed bed. As the fixed bed gets exhausted, the concentration of the adsorbate at the outlet becomes equal to the feed concentration. A plot of exit concentration as a function of contact time or volume throughput treated is known as breakthrough curve. The performance of breakthrough curves is affected by several factors including physical and chemical properties of the adsorbent and adsorbate, column bed height, feed flow rate and pH. Batch experimental data is often difficult to apply directly to fixed bed adsorbers because isotherm cannot give accurate data for scale up since a flow column is not at equilibrium. Therefore experiments in fixed bed were also conducted. The breakthrough plots of Ct/C0 against volume of effluent treated at As(III) and As(V) concentration of 0.5 mg/L at a flow rate of 2 ml/min having different bed heights (6 cm, 12 cm and 18 cm) has been shown in Figures 6 and 7 -0.2 -0.4

As(III) As(V)

log x/m

-0.6 -0.8 -1 -1.2 -1.4 -1.5

-1

-0.5

0

log Ce

Figure 5.

Freundlich adsorption isotherm plot for As(III) and As(V) removal onto AA. 1

Ct/C0

0.8 0.6 0.4

6 cm 12 cm

0.2

18 cm 0 0

10

20 30 40 Volume treated (L)

50

60

Figure 6. Breakthrough plot for As(III) in a continuous mode at different bed height [C0 0.5 mg/L; Flow rate 2 ml/min].

217

1

Ct/C0

0.8 0.6 0.4

6 cm

0.2

12 cm 18 cm

0 0

10

40 50 30 20 Volume treated (L)

60

70

Figure 7. Breakthrough plot for As(V) in a continuous mode at different bed height [C0 0.5 mg/L; Flow rate 2 ml/min]. 1

Ct/C0

0.8 0.6 0.4 As(III) 0.2

As(V)

0 0

10

20

30

40

50

60

70

Volume treated (L)

Figure 8. Comparative study of As(III) and As(V) in a continuous mode [C0 0.5 mg/L; Bed height 18 cm; Flow rate 2 ml/min].

respectively. At lower contact time, the curves were steeper indicating the faster exhaustion of the fixed bed. As can be seen from the breakthrough curves (Fig. 7), increase in treated volume of water was obtained with increased bed height. Examining the adsorption capacities, AA has better As(V) removal efficiency as compared to As(III). A comparative plot showing the faster exhaustion of AA bed in case of As(III) has been shown in Figure 8. Lower bed sorption capacity was observed in case of As(III) compared to As(V) showing a similar behaviour for As(III) removal as in batch system. Main objective of the fixed bed adsorption is to reduce concentration in the effluent so that it does not exceed permissible limit (breakthrough concentration Cb). Bohart & Adams (1920) proposed a relationship between bed depth and time taken for breakthrough to occur. The service time was related to process conditions and operating parameters as (5) Hutchins (1973) proposed a linear relationship between bed depth and service time given by equation (6) where C is the effluent concentration of adsorbate in the liquid phase (mg/L); C0 is initial concentration of sorbate in the liquid phase (mg/L), v is linear flow rate (m3/s), N0 is adsorption capacity (mg solute/g adsorbent), ka is rate constant in BDST model (l/mg h), t is time (s) and Z is bed depth of column (m). Equation (6) enables the service time, t, of an adsorption bed to be determined for 218

Breakthrough time (hr)

300 250 200 150 100

As(III)

50

As(V)

0 5

Figure 9.

10 15 Bed height (cm)

20

BDST plot for As(III) and As(V) [C0 0.5 mg/L; Flow rate 2 ml/min].

a specified bed depth, Z, of adsorbent. The service time and bed depth are correlated with the process parameter such as initial pollutant concentration, solution and flow rate and adsorption capacity. Equation (6) has the form of a straight line (7) where mx is the slope of the bed depth service time (BDST) line and the intercept of this equation represents (8) Bed depth service time plot for As(III) and As(V) adsorption onto AA has been shown in Figure 9 with R2 value ⬎0.99. The slope of the BDST line (Fig. 9), mx, represents the time required for the adsorption zone to travel a unit length through the adsorbent under the selected experimental conditions at a given concentration and in the present study it was found to be 12.2 hours and 15.4 hours for As(III) and As(V) respectively (C0 0.5 mg/L). This is used to predict the performance of the bed, if there is a change in the initial solute concentration, C0, to a new value of solute concentration. The critical bed depth Z0 is obtained for t ⫽ 0 and for a fixed outlet concentration, Ct ⫽ Cb where Cb is the concentration at the breakthrough defined as a limit concentration or a fixed percent of initial concentration. (9) The critical bed depths as obtained from BDST plot for As(III) and As(V) were 0.63 cm and 0.36 cm respectively. The critical bed depth represents (Z0) the theoretical depth of adsorbent, which is necessary to prevent the sorbate concentration to exceed the limit concentration Cb. The data may be utilized for the design of an arsenic removal unit. The exhausted activated alumina was successfully regenerated by using 2% NaOH as regenerant followed by treatment with 1% HCl for reactivation of sorption sites. A reduction of 11% in sorption capacity was observed after five regeneration cycles. Solidifcation/Stablisation studies carried out revealed that AA mixed with ordinary portland cement, fly ash and lime (AA ⫹ C ⫹ FA ⫹ CH) in the weight ratio of 1:1:1:1 can be successfully used for the safe disposal of spent alumina sludge.

4

CONCLUSIONS

Adsorption of arsenite and arsenate on activated alumina was found to be dependent on contact time, pH, adsorbent dose and initial sorbate concentration. Activated alumina was found effective 219

for the As(V) as well as As(III) in the concentration range studied. High sorption capacity of activated alumina could be due to its interior surface properties. Maximum adsorption of arsenite (96%) and arsenate (98.6%) was observed at pH 7.6 and 5.8 respectively. Adsorption of both arsenite and arsenate on activated alumina can be described well with Langmuir and Freundlich isotherms. Both coulombic and specific chemical interaction seems to be involved in the adsorption of arsenate. Kinetic experiments revealed that the adsorption of As(III) and As(V) occurs rapidly in the first 6 and 4 hours respectively and adsorption followed first order rate kinetics. The column experiments show that the adsorption of As(III) and As(V) is a function of bed height, flow rate and contact time. Breakthrough experiments show that As(V) can be more easily removed as compared to As(III). The results of these investigations can be applied for the design of an arsenic removal unit for the treatment of contaminated groundwater.

REFERENCES Bellack, E. 1971. Arsenic removal from potable water. J. American Water Works Association 63: 454. Bohart, G.S. & Adams, E.Q. 1920. Behavior of charcoal towards chlorine. J. Chemical Society 42: 523–529. Bowers, A.R. & Huang, C.P. 1985. Adsorption characteristics of metal-EDTA complexes onto hydrous oxides. J. Colloid Interface Science. 105: 197–215. Cullen, W.R. & Reimer, K.J. 1989. Arsenic speciation in the environment. Chem Rev. 89: 713–764. Dambies, L., Guibel, E. & Rose, A. 2000. As(V) sorption on molybdate impregnated chitosan beads. Colloids And Surfaces-A-Physiochemical Engineering Aspects 170: 19–31. Dutta, A.K., Gupta, A., Bandyopadhyay, P., Ghosh, A., Biswas, R.K., Roy, S.K. & Deb, A.K. 1997. Appropriate technology for removal of arsenic from drinking water of rural West Bengal. Proceedings of Annual Conference of AWWA, Atlanta, Georgea, USA. Elson, C.M., Devies, D.H. & Hayes, E.R. 1980. Removal of arsenic from contaminated drinking water by a chitosan/chitin mixture. Water Research 14(9): 1307–1311. Ferguson, J.F. & Gavis, J.A. 1972. Review of the arsenic cycle in natural waters. Water Research 6: 1259–1274. Ghosh, M.M. & Yuan, J.R. 1987. Adsorption of inorganic arsenic and organ arsenicals on hydrous oxides. Environmental Progress 6(3): 150–157. Gulledge, J.H. & O’Connor, J.T. 1973. Removal of As(V) from water by adsorption on Al and Fe hydroxides. J. American Water Works Association 13: 548–552. Gupta, G.S., Prasad, G., Panday, K.K. & Singh, V.N. 1988. Removal of chrome dye from aqueous solutions by fly ash. Water Air and Soil Pollution 37: 13–24. Gupta, S.K. & Chen, K.Y. 1978. Arsenic removal by adsorption. Journal Water Pollution Control Federation 50: 493–506. Huang, C.P. & Fu, P.L.K. 1984. Treatment of As(V) containing water by the activated carbon process. Journal Water Pollution Control Federation 56: 80–88. Hutchins, R.A. 1973. New method simplifies design of activated carbon system. Chem. Eng. 80: 133–138. Knocke, W.R. & Hemphill, L.H. 1981. Mercury(II) sorption by waste rubber. Water Research 15: 275–282. Lee, J.Y. & Rosehart, R.G. 1972. Arsenic removal by sorption processes from wastewaters. Can. Min. Metallurg B. 65: 33–37. Low, K.S. & Lee, C.K. 1995. Chrome waste as sorbent for the removal of arsenic (V) from aqueous solution. Environmental Technology 16: 65–71. Manju, G.N., Raji, C. & Anirudhan, T.S. 1998. Evaluation of coconut husk carbon for the removal of arsenic from water. Water Research 32: 3062–3070. Morales, K.H., Ryan, L., Kuo, T.L., Wu, M.M. & Chen, C.J. 2000. Risk of internal cancers from arsenic in drinking water. Environmental Health Perspectives 108(7): 655–662. Ng, J.C., Wang, J. & Shraim, A. 2003. A global health problem caused by arsenic from natural sources. Chemosphere 52: 1353–1359. Muzzarelli, R.A.A., Tanfani, F. & Emanuelli, M. 1984. Chelating derivatives of chitosan obtained by reaction with ascorbic acid. Carbohydrate Polymers 4: 137–151. Ohki, A., Nakayachigo, K., Naka, K. & Maeda, S. 1996. Adsorption of inorganic and organic arsenic compounds by aluminum-loaded coral limestone. Applied Organmetallic Chemistry 10: 747–752. Pattanayak, J., Mondal, K., Mathew, S. & Lalvani, S.B. 2000. A parametric evaluation of the removal of As(V) and As(III) by carbon-based adsorbents. Carbon 38: 589–596.

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Pierce, L.M. & Moore, C.B. 1982. Adsorption of arsenite and arsenate on amorphous iron oxide. Water Research 16: 1247–1253. Rosenblum, E. & Clifford, D. 1984. The equilibrium As capacity of activated alumina. (EPA-600/S2-83-107). Singh, D.B., Prasad, G. & Rupainwar, D.C. 1996. Adsorption technique for the treatment of As(V) rich effluents. Colloids Surfaces. A: Physiochemical Engineering Aspects. 111: 49–56. Stumm, W. & Morgan, J.J. 1996. Aquatic chemistry: chemical equilibria and rates in natural waters, (3rd ed) New York, Wiley. Suzuki, T.M., Bomani, J.O., Matsunaga, H. & Yokoyama, T. 1997. Removal of As(III) and As(V) by a porous spherical resin loaded with monoclinic hydrous zirconium oxide. Chemistry Letters 11: 1119–1120. Tokunaga, S., Wasay, S.A. & Park, S.W. 1997. Removal of arsenic (V) ion from aqueous solutions by lan thanum compounds. Water Science Technology 35: 71–78. USEPA 2001. National primary drinking water regulations; arsenic and clarifications to compliance and new source contaminants monitoring; Final Rule. Federal Register. 66(14): 6976-7066. Wasay, S.A., Haron, M.J., Uchimi, A. & Tokunga, S. 1996. Removal of arsenite and arsenate ions from aqueous solution by basic Yttrium carbonate. Water Research 30: 1143–1148.

221

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Comparing the arsenic sorption capacity of Bauxsol™ and its derivatives with other sorbents H. Genç-Fuhrman Environment & Resources, Technical University of Denmark (DTU), Lyngby, Denmark

D. McConchie Southern Cross University, Centre for Coastal Management, East Lismore, N.S.W., Australia

O. Schuiling Institute of Earth Sciences, Utrecht, Netherlands

ABSTRACT: In view of the frequency of occurrence of groundwaters with high arsenic contents in many countries, including several of the poorest developing countries, the development of an inexpensive and efficient system for arsenate removal from drinking water is essential. Treatment of water and effluents with solid industrial residues for adsorptive removal of arsenic offers a potentially costeffective and sustainable approach. Herein, promising sorbents have been developed by the authors from seawater-neutralised bauxite refinery residues (Bauxsol™) including acid treated Bauxsol (ATB), acid and heat treated Bauxsol (AB), Bauxsol coated sand (BCS), and AB coated sand (ABCS). A comparison of the affinity of commonly used adsorbents towards arsenic with that of the sorbents developed by the authors shows that the sorptive capacity of the new materials is comparable to, or better than, that of alternative materials. Because the raw material for the developed sorbents (red mud) is originally an industrial residue, the developed sorbents should be readily available and inexpensive compared to alternative materials and their use will enhance resource use efficiency.

1

INTRODUCTION

Arsenic (As) contamination of drinking water is a major health concern, because drinking arsenic contaminated water, even at quite low concentrations, is linked to several types of cancers. Consequently, increasingly stringent legislation on the permissible concentrations of arsenic in drinking water has led to increased investigations of the occurrence, chemical speciation and mobility of arsenic in natural waters and of methods for removing arsenic during water treatment. Although background arsenic concentrations in natural environments are usually low, arsenic concentrations are high in many parts of the world due to mobilization from natural geological sources or at a smaller scale from industrial pollution. Arsenic, associated with industrial pollution, can be managed by improving process engineering and environmental management practices; however making water that has naturally high arsenic content safe to drink requires some form of water treatment to reduce arsenic concentrations. Unfortunately, there is no known cure for arsenic poisoning and therefore providing arsenic free drinking water is the only way to diminish the adverse health affects of arsenic. Several methods are proposed to provide arsenic-free water. These methods suggest either the treatment of arsenic contaminated groundwater, or investigating alternative sources (e.g. surface water treatment, rain-water harvesting, etc). The use of alternative water sources, however, can only be possible after a major and costly technological shift and thus, the treatment of arsenic contaminated water to the guideline values is the preferred option (Ahmed 2003). High concentrations of arsenic in water and soil have been documented in Taiwan, Argentina, the USA, Chile, Hungary, China, Greece, Mexico, Vietnam, 223

Ghana, Thailand, Canada, Germany and many other countries, but of these, the most severe outbreaks of arsenic poisoning have been associated with groundwaters in the Bengal Delta area including Bangladesh and West Bengal (Eastern India), where it has been estimated that as many as 120 million people (80 million in Bangladesh, 40 million in India) are at risk (Chowdhury et al. 2000). Arsenic contamination in the affected districts of the Bengal Delta is potentially a major environmental disaster (Karim 2000, see Ahmed et al. 2004 for an extensive review). Arsenic is also a social concern in Bangladesh because women affected by arsenic are reportedly discriminated against in their working environments and many have to leave their jobs when skin changes caused by arsenic contaminated water become apparent, leading to economic hardship and social disruption (The Daily Star 2003). Arsenic has a rather complicated chemistry, because it can exist in several forms in the environment, but dominant forms of arsenic in natural waters are, inorganic As(III) or arsenite and inorganic As(V) or arsenate. Both arsenite and arsenate have a high affinity for Fe-oxides (Goldberg & Johnston 2001, Smedley & Kinniburgh 2002), but the cost of the adsorptive metal removal process is high when pure sorbents (either activated carbon or hydrated Fe- and Al-oxides) are used (Apak et al. 1998). Consequently, the cost of pure adsorbents may be a limitation for many water treatment applications and there is a strong motivation to find cost-efficient alternatives; e.g. red mud, which is used in the present study after neutralization with seawater. Red mud is a fine-textured insoluble residue remaining after high temperature caustic digestion when the Bayer process is used to extract alumina (Al2O3) from bauxite (Chvedov et al. 2001). In 2003, roughly 22 million tones of alumina were produced worldwide and 10% of this was in Asia. This translates to an annual production of roughly 35 million tones of red mud residue worldwide, of which about 3 million are produced in Asia (WAO 2003), where the arsenic contamination of drinking water is particularly serious. In short, red mud is widely available in large quantities, it is rich in Fe- and Al-oxides/hydroxides (Altundog˘an et al. 2000), and it is expected to have a strong affinity for arsenic. Currently, most red mud is stored in ponds or dams (López et al. 1998), although some refineries dump their red mud in the sea. The red mud dams occupy a large area and can constitute a serious environmental hazard due to the highly caustic nature of the red mud, the fact that it dries exceptionally slowly, and the fact that wet red mud has a very low physical strength (Hind et al. 1999). The stored red mud can pollute environmental compartments such as soil, surface water and ground water, and it can pose a serious threat to lives and properties if a red mud dam bursts (Varnavas & Achilleopoulos 1995, Zhang et al. 2001). Thus, the reuse of red mud residues is a potentially costeffective alternative to long-term storage in impoundments (McConchie et al. 1999, 2002a, 2002b). Red mud is thus of interest to researchers for many reasons and it has been the subject of many detailed investigations. The particular interest of the authors was the assessment of potential uses for seawater-neutralised red mud (Bauxsol™) for water treatment, particularly in relation to arsenic removal, both to provide an economically viable water treatment processes and for environmental protection. With this motivation Bauxsol and Bauxsol derivatives such as acid treated Bauxsol (ATB), acid and heat treated Bauxsol (also called Activated Bauxsol (AB)), ferric or aluminum sulfate supplemented Bauxsol and AB, Bauxsol coated sand (BCS) and AB coated sand (ABCS) have been previously developed by the authors and used to remove arsenic from water. Details of these studies are reported in Genç et al. (2003), Genç & Tjell (2003) and Genç-Fuhrman et al. (2004a, 2004b, 2004c, 2004d). The ultimate objective of the present study is to compare the sorptive capacity of the previously developed Bauxsol and Bauxsol derivatives with the other arsenic sorbents and to investigate the potential of using Bauxsol as a novel adsorbent for arsenic removal from water.

2

BAUXSOL AND BAUXSOL DERIVATIES

Five different sorbents, previously developed by the authors, namely Bauxsol, ATB, AB, BCS, and ABCS are compared with other arsenic sorbents, which have been reported in the literature, in relation to their arsenic removal capacity determined using batch and column studies. A simple sorbent production scheme is presented in Genç-Fuhrman (2004d), and is represented here in Figure 1. 224

225

Figure 1. Preparation of seawater-neutralised red mud (Bauxsol), acid treated Bauxsol (ATB), Activated Bauxsol (AB), Bauxsol coated sand (BCS), AB coated sand (ABCS), and ferric or aluminum sulfate added Bauxsol and AB.

Details on the preparation of each sorbent can be found in Genç et al. (2003) and Genç-Fuhrman et al. (2004a, 2004d). 2.1

Seawater-neutralised red mud (Bauxsol)

Untreated red mud is highly caustic (the pH is usually ⬎13.0), due to the presence of residual NaOH remaining after bauxite digestion and the formation of some Na2CO3. This high basicity is environmentally hazardous, and the red mud needs to be neutralised before it can be used for water treatment; it also needs to be neutralised if it is to be stored safely or regenerated. Several methods for the neutralisation of red mud have been previously reported e.g. involving the addition of gypsum or acid (López et al. 1998) or using either Ca- and Mg-rich brines or water with added CaCl2 and MgCl2 (McConchie et al. 2002a). The Bauxsol and its derivatives reported in this study have been seawater-neutralised (McConchie et al. 1999, 2002a), as seawater-neutralisation is both cheap and simple. The composition and mineralogy of major components in the Bauxsol used in this study is reported in Genç et al. (2003), and additional compositional and mineralogical details can be found elsewhere (McConchie et al. 1999, Clark 2000). The chemistry of seawater-neutralisation is described in detail by McConchie et al. (1999), but essentially during the neutralisation process hydroxyl ions in the red mud are neutralised largely by reaction with magnesium to form brucite and hydrotalcite, with calcium to form hydrocalumite and p-aluminohydrocalcite; and carbonate ions react with calcium in the seawater to form calcite and aragonite. 2.2 Acid treated Bauxsol (ATB) and activated Bauxsol (AB) Acid treatment, and acid plus heat treatment of Bauxsol is carried out as suggested by Pratt & Christoverson (1982) to improve the sorptive capacity of Bauxsol. During acid treatment with hydrochloric acid (HCl), sodium compounds that can adversely affect arsenic adsorption are removed (Shiao & Akashi 1977, Altundog˘an et al. 2002, Genç-Fuhrman et al. 2004a). Here, HCl is selected for the acid treatment over sulphuric acid (H2SO4), because HCl can yield significantly more Fe(H2O) 63⫹ and soluble hydroxy-complexes than can H2SO4 (Apak & Ünseren 1987). Moreover, the addition of sulfate to the system is undesirable, as sulfate diminishes the arsenate removal efficiency (Genç & Tjell 2003, Genç-Fuhrman et al. 2004a). When the sorptive capacities of Bauxsol and the sorbents obtained from acid treatment and the combined acid and heat treatment are compared, it is found that the combined acid and heat treated Bauxsol (AB) performs the best, indicating further improvement of the sorptive capacity during the heat treatment. The chemical and mineralogical characteristics of Bauxsol and AB are presented in Table 1, where it can be seen that the Fe- and Al-oxide content (note that Fe- and Al-oxides possibly contribute for most of the arsenic removal when using Bauxsol or its derivatives) of AB is greater than that of Bauxsol. After the treatment, a roughly 300% increase in surface area and Table 1. Chemical compositions of Bauxsol and activated Bauxsol (AB) (after Genç-Fuhrman et al. 2004b). Chemical composition % (w/w)

% (w/w)

Constituent

Bauxsol

AB

Constituent

Bauxsol

AB

Fe2O3 Al2O3 SiO2 TiO2 CaO Misc.

34.05 25.45 17.06 4.90 3.69 9.86

46.55 26.51 17.4 6.9 0.7 –

Na2O MgO K2O P2O5 MnO

2.74 1.86 0.20 0.15 0.04

0.5 0.5 0.4 0.4 0.1

226

cation exchange capacity is observed for AB (Genç-Fuhrman et al. 2004a), which may partly account for the increase in the adsorption capacity. The removal of sodium could also account for much of the increased surface area (Pratt & Christoverson 1982). Most importantly, the arsenate removal capacity is significantly increased by the treatment (Genç-Fuhrman et al. 2004a). It is, however, important to note that although the surface area of AB is roughly two times greater than that of ATB, the arsenate removal obtained using AB and ATB are 100% and 95%, respectively. This observation may reflect the fact that the micropores, which primarily contribute to the surface area, did not actively take part in the adsorption (Elizalde-Gonzáles et al. 2001), as the observed difference for the arsenate removal is less than the relative difference in surface area. Another difference between the Bauxsol and AB, that could affect arsenate sorption, is that unlike slightly alkaline Bauxsol, AB has a near-neutral reaction pH when suspended in water. 2.3

Ferric- or aluminum sulfate added Bauxsol and AB

In addition to applying activation methods described above, another attempt was made to increase the arsenic removal efficiency by adding ferric sulfate (Fe2(SO4)3.7H2O) or aluminum sulfate (Al2(SO4)3.18H2O) to the Bauxsol and AB. As expected, the addition of these reagents reduced the pH to 4.5, due to hydrolysis reactions (e.g. Fe(III) ⫹ 3H2O → Fe(OH)3) (Sutherland et al. 2002). Nevertheless, the addition of these reagents did not enhance arsenic removal and in fact, their addition reduced arsenate sorption slightly. Aluminum sulfate most significantly suppressed arsenate removal, whereas the suppressive effect of ferric sulfate was minor. Later experiments were repeated using AB and similar results were obtained for AB following the addition of ferric or aluminum sulfate (Genç-Fuhrman et al. 2004a). Several reasons are postulated to explain these observations, including sulfate competition for the available adsorption sites, formation of watery gels covering the surface, and unfavorable pH conditions (Genç-Fuhrman et al. 2004a). With the benefit of hindsight, it seems likely that better arsenic removal would have been obtained by using ferric chloride instead of ferric or aluminum sulfate; because both sulfate competition and gel formation would have been greatly reduced. However, Ruhland & Jekel (2002) report that the ferric chloride backwash sludge may be less desirable because ferric sulfate sludge has better characteristics for sedimentation and thickening. Clearly, further investigations will be required to resolve these questions. 2.4

Bauxsol coated sand (BCS) and AB coated sand (ABCS)

Because Bauxsol and AB contain very fine particles, they are not suitable for use as filtration media in fixed bed column tests where they may cause severe clogging. Thus, Bauxsol and AB are coated to sand as described in Genç-Fuhrman et al. (2004d) for use in column tests, as well as in batch tests to obtain the adsorption capacity (see Fig. 1 for a simple preparation scheme).

3

RESULTS AND DISCUSSIONS

Bauxsol and its derivatives are compared with some other widely used arsenic sorbents reported in the literature in terms of their adsorption capacity. Adsorption data in the literature cannot, however, be easily compared due to operational differences in experimental procedures (Xu et al. 1991). Thus, comparisons are made by reporting the experimental conditions along with the adsorption capacity data. When the affinity of commonly used adsorbents towards arsenic is compared with Bauxsol and its derivatives (see Table 2) it is found that the obtained sorptive capacities for Bauxsol and its derivatives are comparable to, or better than, that of other available sorbents. It can be seen from Table 2 that arsenate adsorption is generally favored at low pH values, whereas for arsenite high pH values are preferred. At the given experimental conditions, Bauxsol and its derivatives can effectively remove arsenic (mainly arsenate) from water down to acceptably low concentrations but, their respective 227

228

Activated Alumina Activated Alumina Activated Alumina Activated Bauxite Activated Carbon Activated Carbon Activated Red Mud Activated Bauxsol (AB) Activated Bauxsol (AB) e ABCS e ABCS f Al-SZP1 Bauxsol g BCS g BCS g BCS Char Carbon h CMFB Feldspar i Fe-LDA j Fex(OH)y-Montm.

8.3 6.9 – 7.3 – – 7.25 6.8 – – – – – – – – 3.1 – – 8–10 9

Batch Batch Batch Batch Batch Batch Batch Batch Batch Batch Column Batch Batch Batch Batch Column Batch Batch Batch Batch Batch

Sorbents

Ce, mg/L

b

6.5 5.2 7.7 6.4 4.0 3.2 2–3.5 4.5 7.1 7.1 7.4 3–10 6.3 7.1 4.5 5.7 3.0 3 4.2 2–4 9

0.2 15.9 – 0.1 – – 0.89 0.5 – – – – – – – – 89.2 – – 43.7 13

4.1 3.5 1.1 2.0 3.5 0.3 0.94 7.7 3.0 2.1 1.6 4.9 1.1 1.6 3.3 6.2 34.5 57.9 0.2 38.5 4

0.05 4.03 – 0.05 – – 28.65 11.25 – – – – – – – – 536.7 – – 0.30 45

0.05 7.53 0.063 0.05 0.01 0.05 28.58 0.03 0.26 5 5 0.01 0.05 5.5 4 1 172.3 200 – 0.19 55

As(V) As(III) As(V) As(III) As(V)

Adsorption maxima, mg/g

2 0.1–0.5 0.5–5 2 1 3 20 5 5 5 140 5 10 5 5 140 5 1 40 50 1.6

X, g/L

c

Comparing several sorbents reported in the literature for arsenic removal at the ambient temperature.

Experimental apH Set-up, Batch/Column As(III)

Table 2.

Gupta & Chen (1978) Lin & Wu (2001) Wang et al. (2002) Gupta & Chen (1978) Huang & Fu (1984) Gupta & Chen (1978) Altundog˘ an et al. (2002) Genç-Fuhrman et al. (2004b) Genç-Fuhrman et al. (2004b) Genç-Fuhrman et al. (2004d) Genç-Fuhrman et al. (2004d) Xu et al. (1998), Xu et al. (2002) Genç et al. (2003) Genç-Fuhrman et al. (2004d) Genç-Fuhrman et al. (2004d) Genç-Fuhrman et al. (2004d) Pattanayak et al. (2000) Loukidou et al. (2003) Singh et al. (1996) Matsunaga et al. (1996) Lenoble et al. (2002)

References

Commercially available Commercially available Commercially available No information is given – 2 1 – – 2 1 1 No information is given – 2 1 2 – 2 – 1 1 – – 1 – 1 – 1 1 – – 1 – 1 – – 1 – – 1 – 1 – – 1 – – 1 – 1 – – – 2 – 2 2 1 1 1 1 – 1 – – 2 – 1 1 – – – – 2 – – 1 – – – – 2 – 1 1 – – – – 2 – 1 – 2 – 2 2 2 – 1 – 2 – 2 – 2 – 1 1 – 1 – 1 2 – 1 – 2 – 2 – 2 – 1 – 2 – 2 – 2 – 1

1 1 – – – 1 1 – 1 1 1 1 1 1 1

1 – 1 1 – – 1 1 1 1 – – 1 1 –

2 – – – – 2 – – – – 2 – 1 1 2

1 1 –

a b c d e f g h i j k

Cost parameters

d

229

Batch Batch Batch Batch Column Batch Batch Batch Column Batch Batch Column Batch Column Column Batch Batch Batch Batch Batch Batch Column Batch Batch

– – 6.3 9 – – 7.6 – 7.6 – 9 7.2 7.6 7.6 3 – 9.5 – – 9 – 6.5–7.2 6.5 4.0

7.0 4.0 6.5 9 7.8 7.8 7.6 4.2 7.6 7.0 9 7.2 7.6 7.6 3 7.0 3.2 7.5 3–3.3 9 7 – 4.0 4.0

– – 0.5 22 – – 0.1 – 0.8 – 28 – 0.04 0.4 p 80 – 0.67 – – 13 – 4.4 31.7 0.02

0.2 27.8 15.4 4 8.5 8.5 0.2 0.2 0.9 m 14 7 – 0.04 0.3 p 230 0.2 0.52 93.1 47 2 – – 30.7 0.1

– – 0.016 30 – – 0.05 – 0.05 – 20 0.05 0.05 0.005 450 – 16.7 – – 47 – ⬎0.02 ⬍0.01 0.01

0.5 – ⬍0.002 57 0.01 0.01 0.05 0.11 0.05 0.07 30 0.05 0.05 0.005 450 450 19.7 25 0.01 56 20–200 – ⬍0.01 0.01

5 1 0.4 1.6 – l 1 5 40 – n 0.014 1.6 – – – – 10 20 0.5 1 1.6 2.5 74.3 – y 110

Bhattacharyya et al. (2003) Diamadopoulos et al. (1993) Chakravarty et al. (2002) Lenoble et al. (2002) Driehaus et al. (1998) Driehaus et al. (1998) Thirunavukkarasu et al. (2003) Singh et al. (1996) Thirunavukkarasu et al. (2003) Holm, (2002) Lenoble et al. (2002) DeMarco et al. (2003) Thirunavukkarasu et al. (2003) Thirunavukkarasu et al. (2003) Dambies et al. (2002) Ouvrard et al. (2001) Altundog˘ an et al. (2000) Deliyanni et al. (2003) Matis et al. (1997) Lenoble et al. (2002) Ramaswami et al. (2001) Nikolaidis et al. (2003) Suzuki et al. (2000) Elizalde-Gonzáles et al. (2001)

1 – 1 – – 2 – 1 1 – 1 – 1 2 2 1 1 – 1 – – 1 3 1 – 2 – 2 – 2 – 1 – 2 – 2 1 2 3 1 – 2 – 2 1 2 3 1 – 2 – 2 – 2 – – 1 – 1 – 1 2 – 1 – – – – 1 1 – 1 No information is given – 2 – 2 – 2 – 1 – 2 – 2 2 – 3 – – 2 2 – – – – – – 2 2 – – – – – – 2 – 2 1 2 – 1 – 2 – 2 1 2 1 1 1 – 1 – 1 2 3 1 – 2 – 2 1 2 2 – 1 – – 2 – 2 1 0 – 2 – 2 – 2 – 1 Commercially available Commercially available – 2 – 2 2 1 – 1 – 2 – 2 1 2 – 1

– – – – – 0 1 1 1 –

1 – – – 1 1 1 – 1 1

e

2 – – – 2 2 2 2 – 2

1 2 – 2 – – – 1 –

1 1 – 1 – –

– 1 – – 1 1 – 1 2

1 1 – 1 1 1 – 1 1

Note: –, Data not available or not relevant, a optimum pH value, bequilibrium sorbate concentration, c in case of column tests the weight of the filter medium is given in g, see Table 3, f tap water is used, g aluminum-loaded Shirasu-zeolite, h chemically modified fungal biomass (P. chrysogenum), i iron(III)-loaded chelating resin with lysine- N␣, N␣-diacetic acid, j montmorillonite pillared with titanium(IV), k ferruginous manganese ore, l given as gFe/L, m mg/gFe, n mgFe/L o hybrid ion exchanger, p given as mg As/g Mo, r ␤-FeO(OH), s montmorillonite pillared with iron(III), t field experiments, u zirconium(IV) loaded polymer resin functionalized with diethlenetriamine-N,N,N’,N’polyacetic acid, v clinoptilotile-rich Mexican tuff, y given as zeolite in a mass/L.

Ferralite Fly ash k FMO Goethite Granular ferric hydroxide Granular ferric hydroxide Granular ferric hydroxide Hematite (␣Fe2O3) Granular ferric hydroxide HFO HFO o HIX Iron oxide coated sand Iron oxide coated sand Modified chitosan gel Natural Manganese Oxide Red Mud r Synthetic Akaganeite Synthetic goethite s TixHy-Montm Zero-valent iron t Zero-valent iron u Zr-CMA resin v ZMS

affinities towards arsenic, as well as their optimum operational conditions differ. The affinity of the developed sorbents towards arsenic in a decreasing order is AB ⬎ ATB ⬎ ABCS ⬎ BCS ⬎ Bauxsol, and sorptive capacities of all tested sorbents compare well with those of similar unconventional sorbents. Moreover, AB, ATB and ABCS can even compete well with the pure sorbents such as Fe- and Al- oxides commonly employed for arsenic removal and with conventional precipitation or flocculation methods. The arsenic removal process, when using Bauxsol or its derivatives, is probably not simple adsorption; and the process probably involves more than one arsenic removal mechanism and this may be a particular advantage. However, it is still not fully understood how the removal works and it may not be possible to determine the relative contributions of simple adsorption, absorption, adsorption precipitation and solid state diffusion (Apak et al. 1999). Sorption is the common term used for adsorption, surface precipitation or absorption (Sposito 1986), and it is used in this broad sense here when referring to overall arsenic removal using the developed sorbents. The sorption of arsenic using Bauxsol or AB possibly takes place by three mechanisms: (i) formation of surface precipitates; (ii) co-precipitation (with diffusion or dissolution); and (iii) adsorption (Krauskopf & Bird 1995). For the sake of simplicity, the initial removal process is assumed to be adsorption, and the adsorption data studied using the Langmuir and Freundlich isotherms. However, it should be kept in mind that subsequent reactions, such as surface precipitation, solid solution formation and diffusion into the absorbent are also possible in the system. The cost of the pure sorbents that work well for arsenic removal is a major limitation, and the sorbents developed in this study are cost-effective, as the raw material for the developed sorbents, red mud, is originally an industrial residue. Moreover, the developed sorbents should be readily available and inexpensive compared to alternative materials and their use will enhance resource use efficiency. The use of the sorbents developed during this study may also produce several other benefits over similar sorbents, because these sorbents will simultaneously remove a wide range of other potentially hazardous trace elements to very low concentrations without introducing any other secondary contaminants to the water (McConchie et al. 1999, 2002a). These additional benefits will be achieved at no extra cost and because the spent sorbents are not toxic, it is postulated that they can be disposed of without the need for confinement. The numerous tests, i.e. the toxicity characteristic leaching procedure (TCLP) test (USEPA 1996), already conducted indicate that even the spent Bauxsol reagents will pass the standard leaching tests for classification as inert solids such that they don’t need to be disposed of in toxic waste facilities (Genç et al. 2003, Genç-Fuhrman et al. 2004a). But a detailed study on the disposal method of spent Bauxsol, AB, BCS and ABCS should still be carried out. Furthermore, the developed sorbents need to be tested using natural water samples before drawing final conclusions and full-scale studies need to be carried out, because laboratory-scale tests may provide only an

Table 3.

Cost unit-criteria matrix (used to estimate the cost of the sorbents that are compared in Table 2).

Cost unit

Criteria

1 2 1 1⫹1 3

a) b) c) d) e)

2 3 1 1 1 2

f) g) h) i) j) k)

Adsorbent prepared in laboratory using a cheap method Adsorbent prepared in laboratory using an expensive method (or purchased) Preparation of the adsorbent requires minimum men power and energy Preparation of the adsorbent requires skilled men power and significant energy Prepared adsorbent is toxic and needs to be handled with care during storage, transportation, and etc. The raw material needs to be transported Spent adsorbent is toxic and needs to be disposed with special care pH needs to be adjusted before the process pH needs to be adjusted after the process Pre-oxidation of As(III) to As(V) is necessary The guideline value of 0.01 mg/L can be not reached, additional treatment is necessary

230

approximation of processes at pilot- or full-scale levels (Cheng et al. 1994, Scott et al. 1995, Hering et al. 1997). Note that lower efficiencies are often reported in full-scale studies compared to those obtained in the laboratory (McNeill & Edwards 1995). Furthermore, if the sorbents are to be tested in Bangladesh and India, where traditions and religious beliefs are very important in daily life, the social acceptance of the sorbent by the local population should also be evaluated because the sorbent is originally a waste material.

4

APPLICABILITY OF THE METHOD

In this section the usefulness of Bauxsol, ATB, AB, BCS and ABCS to remove arsenic from drinking water is considered in relation to the likely cost of the sorbents. Red mud is the raw material of the sorbents used in this study and the (qualitative) cost of preparing Bauxsol, acid treated Bauxsol, AB, BCS and ABCS from red mud can be summarized as follows: 1. Supplying red mud: Red mud can be obtained free from alumina refineries, but transportation of the material may be a concern. Note that alumina refineries are located worldwide, including in areas geographically close to the arsenic contaminated areas (e.g. India). Hence, the transportation cost may not be a serious impediment. For example, the transportation costs of the sorbents from the available areas to the arsenic affected areas might be estimated as US$16/ton/1000 km, as postulated by Bhattacharyya et al. (2003), for the transportation of ferralite in Bangladesh. This figure is used here as an initial approximation. 2. Seawater-neutralisation of red mud: The main costs for this process are for managing the seawater itself, as well as for the basic labor for mixing and subsequently separating the solid and liquid phases. If the neutralised red mud is to be washed, or dried, or both, these additional requirements will increase costs. 3. pH adjustment: The reaction pH of Bauxsol (fully neutralised) is between 8.2–8.8, which provides the optimum pH for arsenite sorption. Unfortunately, optimum arsenate removal requires a lower pH and thus, a chemical (e.g. HCl) for reducing the pH, and a pH meter to measure the pH should be included in the cost. 4. Acid treatment: Acid treatment may be carried out by boiling HCl and Bauxsol, which requires energy as well as skilled operators to carry out the boiling and the following precipitation processes safely. 5. Heat treatment: Here, a substantial energy input and a suitable kiln or oven is required to carry out the treatment although skilled operators are not required. 6. Coating sand with Bauxsol or AB: Sand and basic labor are needed. After coating BCS and ABCS may be dried either in a kiln or oven or simply under the sun (where possible). Pertinent cost parameters for the preparation of Bauxsol and the other sorbents derived from Bauxsol are given in Table 4; the cost parameter numbers are those presented in the preceding Table 4. Cost parameters for seawater-neutralised red mud (Bauxsol) and Bauxsol based sorbents. Cost parameters Sorbents

1

2

3

4

5

6

Bauxsol ATB AB BCS ABCS

⫹ ⫹ ⫹ ⫹ ⫹

⫹ ⫹ ⫹ ⫹ ⫹

⫹ ⫹ ⫺ ⫹ ⫺

⫺ ⫹ ⫹ ⫺ ⫹

⫺ ⫺ ⫹ ⫺ ⫹

⫺ ⫺ ⫺ ⫹ ⫹

⫹: Cost parameter is relevant, ⫺: cost parameter is irrelevant.

231

numbered list. The preparation of the sorbents is generally not expensive, although the process may be time consuming and it is unlikely that any of the key steps could be carried out at a local household scale. However, Bauxsol and BCS are easy to prepare and perhaps the best approach would be to prepare the sorbents in a commercial-scale facility and supply them to village-scale operators or households for the preparation and filling of simple flow-through filtration systems. Later more advanced, and possibly cheaper, alternatives may be developed, such as the highly porous Bauxsol pellets recently developed by Virotec International, but this remain a topic for further research (McConchie 2004). In relation to the economics of the use of Bauxsol derivatives, it is noted that there is not a great difference in arsenic binding capacity between the acid treated Bauxsol (ATB) and the combined acid and heat treated Bauxsol (AB); i.e. 95 and 100% removal for ATB and AB, respectively. Thus, in the future a more detailed study could be carried out to investigate the sorption characteristics of ATB, and to evaluate whether it is more cost-effective to use simple acid treated Bauxsol compared to AB. Moreover, new blends may be prepared using mixtures of Bauxsol and AB or other similar combinations. There are many possibilities that can be developed to suit each specific application, depending on the source water composition, pH, initial arsenate concentration, etc. Although the TCLP test and other leaching tests show that the spent Bauxsol and AB are not toxic (Genç et al. 2003, Genç-Fuhrman et al. 2004a), the disposal of spent Bauxsol, AB, BCS and ABCS remains to be addressed. This is an inherent issue with all arsenic removal technologies, because the removed arsenic and the spent sorbent must be placed elsewhere (Ramaswami et al. 2001). The method requires further testing in the field to assess the validity of experimental results obtained in the laboratory for arsenic removal using the developed sorbents. This is necessary, because somewhat lower efficiencies are reported in full-scale treatment plants (Scott et al. 1995, Johnston & Heijnen 2001). Overall, the findings of the previous studies and this study indicate that in an arsenic removal plant where Bauxsol, AB, BCS or ABCS are to be used, arsenic removal may be accomplished by employing the following treatment units: (i) Pre-oxidation (to oxidize arsenite to arsenate); (ii) sorption in a batch or column system (to remove soluble arsenic), and (iii) filtration (to remove particulate arsenic).

ACKNOWLEDGEMENTS The authors would like to thank “Virotec International Ltd., Australia” for partially financing this study. Jens C. Tjell from Environment & Resources DTU is also acknowledged for valuable discussions.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Optimization of the removal of arsenic from groundwater using ion exchange C.N. Mulligan, A.K.M. Saiduzzaman & J. Hadjinicolaou Department of Building, Civil and Environmental Engineering, Concordia University, Montreal, Quebec, Canada

ABSTRACT: Various technologies have been evaluated for removal of arsenic from the groundwater. Ion exchange using anion exchange resins is used primarily for water treatment with low amounts of sulfate. Oxidation of this form to As(V) is required before removal. In addition, the production and disposal of spent regenerants and media has also not been studied extensively. Although residual production needs to be minimized as much as possible, there is inadequate data regarding the amounts and compositions of the residuals generated by ion exchange processes and methods for disposal. This study involves the optimization of the ion exchange process and the conditions for regeneration of the ion exchange media. Removal efficiency increased as the pH decreased from 8.5 to 6.5, concentration increased from 500 to 1600 ppb and velocity decreased from 1.94 cm/min to 0.65 cm/min. Regeneration of the ion exchange media enabled desorption of 97% of the adsorbed arsenic.

1 1.1

INTRODUCTION Arsenic groundwater contamination

The use of groundwater as a source of potable water supply presents serious problems in many countries including Bangladesh, India, Thailand, Japan, China, US, and Canada, amongst others. Until recently, it was believed that in most cases, the groundwater is pure enough for drinking, while the surface waters must be treated (purified) before drinking. Therefore, groundwater is preferred. The existence of arsenic and other groundwater issues in these countries have changed things dramatically. All through the past three decades there was a move forward to reduce the serious mortality, predominantly among infants, caused by diseases in surface waters in Bangladesh and other countries. International aid agencies such as UNICEF became involved in funding the drilling of shallow tubewells to gain access to groundwater for domestic supply, which was, uncontaminated by bacteria and otherwise believed to be clean. More than 10 million wells were constructed and it was greatly successful in reducing child death in Bangladesh. Although arsenic is a widely distributed element in the earth’s crust, but it was not generally found in a water-soluble form and thus does not cause a risk to the safety of drinking water supplies (Kinniburgh & Smedley 2001). Arsenic problems have long been recognized to occur in sulphide-rich metaliferous strata (principally, therefore, in particular mining areas) and in some geothermal areas. There were also reports in the literature of arsenic occurring in some arid or semi-arid inland basins, for example in parts of Argentina and the United States but the existence of arsenic in soluble form in the anaerobic groundwater of alluvial and deltaic plains was not by and large recognized until 1995. Arsenic containing minerals include arsenopyrite (FeAs), realgar (AsS), orpiment (As2S3), niccolite (NiAs) and cobalite (CoAsS) (Boyle & Jonasson 1973). Arsenic is a contaminant that originates from and is transported to natural waters through erosion and dissolution of arsenic-containing rocks and soil. Climate and redox potential are significant factors in the transport of arsenic. Arsenic sulfides can be oxidized, releasing arsenic to the environment. 237

Arsenic occurs in natural waters in both organic and inorganic forms but the inorganic arsenic is the most prevalent and is the most likely to exist in significant concentrations. The valence and species of inorganic arsenic are dependent on the oxidation–reduction conditions and the pH of the water. Environmental arsenic is mainly found in two forms: arsenic V (arsenate), and arsenic III (arsenite). Arsenic V is the common oxidized state found in surface water and some groundwater sources. Arsenic III is un-oxidized and mainly found in deep anaerobic groundwater sources. 1.2 Treatment technologies Due to the problems of arsenic in the groundwater, economic solutions need to be found to ensure the safety of the drinking water. Several common treatment technologies are used for removal of inorganic contaminants, including arsenic, from drinking water supplies. Large-scale treatment facilities often use conventional coagulation with alum or iron salts followed by filtration to remove arsenic. Lime softening and iron removal also are common, conventional treatment processes that can potentially remove arsenic from source waters. In small communities small-scale systems often use ion exchange adsorption because of their ease of handling and sludge-free operations (Clifford 1999). Treatment options identified by EPA include ion exchange, reverse osmosis, activated alumina, nanofiltration, electrodialysis reversal, coagulation/filtration, lime softening, greensand filtration and other iron/manganese removal processes, and emerging technologies not yet identified (USEPA 2003). Pre-oxidation technology includes chlorination, potassium permanganate, and ozone, although aeration over a significant time period is also possible (Wang et al. 2000). Both chlorine and potassium permanganate can oxidize arsenite to arsenate within a minute (pH 6.3 to 8.3). Chlorine is low cost but can lead to formation of disinfection products, whereas permanganate is not a disinfectant does not form byproducts, can be regenerated but is more costly. Ion Exchange (IX) can effectively remove arsenic using anion exchange resins. It is recommended as a BAT (best available technology) primarily for sites with low sulfate because sulfate is preferred over arsenic. Sulfate will compete for binding sites resulting in shorter run lengths (USEPA 2003). Nitrate and nitrite may cause some problems since they will also adsorb onto the fresh media and then are desorbed by arsenic. These increased levels could be hazardous. Therefore, ion exchange is recommended for sulfate levels less than 50 mg/L and nitrite and nitrate levels less than 5 mg/L. Total dissolved solids should also be less than 500 mg/L. Activated Alumina (AA) is an effective arsenic removal technology; however, the capacity of activated alumina to remove arsenic is very pH sensitive. It is limited to a pH range 5.5 to 6. High removals can be achieved over a broad range of pH, but shorter run lengths will be observed at higher pH. Because arsenic is strongly adsorbed to the media, only about 50–70% of the adsorbed arsenic is removed. The brine stream produced by the regeneration process then requires disposal (USEPA 2003). Reverse Osmosis (RO) can provide arsenic removal efficiencies of greater than 95% when the operating pressure is ideal. Water rejection (on the order of 20–25%) may be an issue in waterscarce regions and may prompt systems employing RO to seek greater levels of water recovery. Water recovery is the volume of drinking water produced by the process divided by the influent stream (product water/influent stream). Increased water recovery is often more expensive, since it can involve recycling of water through treatment units to allow more efficient separation of solids from water. This can also produce more concentrated solid wastes. However, the waste stream will generally not be as concentrated as anion exchange brines, so it should be easier to dispose of. Although reverse osmosis is listed as a BAT as it removes over 95% arsenic, it was not used to develop national costs because other options are more cost effective and have much smaller waste streams (USEPA 2003). It could however, be cost effective if it is utilized for multipurposes such as removal of total dissolved solids (TDS) and arsenic. Modified Coagulation/Filtration (C/F) is an effective existing treatment process for removal of As(V) according to laboratory, pilot-plant, and full-scale tests. The type of coagulant and dosage used affects the efficiency of the process. Below a pH of approximately 7, removals with alum or ferric sulfate/chloride are similar. Above a pH of 7, removals with alum decrease dramatically 238

(at a pH of 7.8, alum removal efficiency is about 40%). Other coagulants are also less effective than ferric sulfate/chloride. Systems may need to lower the pH or add more coagulant to achieve higher removals over 90% (USEPA 2003). Enhanced Lime Softening (ELS), operates optimally at a greater than pH 10.5 is likely to provide a high percentage of arsenic removal. Systems operating lime softening at lower pH values will need to increase the pH to achieve higher removals of arsenic (USEPA 2003). Magnesium addition may also be required. Sludge production is increased compared to lime softening if arsenic removal is required. Electrodialysis Reversal (EDR) can produce effluent water quality comparable to reverse osmosis. EDR systems are fully automated, require little operator attention, and do not require chemical addition. EDR systems, however, are typically more expensive than nanofiltration and reverse osmosis systems. It should be noted that while electrodialysis reversal is listed as a BAT, it was not used to develop national costs because other options are more cost effective and have much smaller waste streams (USEPA 2003). Oxidation/Filtration (including greensand filtration) has an advantage in that there is not as much competition with other ions. Arsenic is co-precipitated with the iron during iron removal. Sufficient iron needs to be present to achieve high arsenic removals. One study recommended a 20:1 iron to arsenic ratio. Removals of approximately 80% were achieved when iron to arsenic ratio was 20:1. When the iron to arsenic ratio was lowered (7:1), removals decreased below 50%. The presence of iron in the source water is critical for arsenic removal. If the source water does not contain iron, oxidizing and filtering the water will not remove arsenic. When the arsenic is present as As(III), sufficient contact time needs to be provided to convert the As(III) to As(V) for removal by the oxidation/filtration process (USEPA 2003). The removal of arsenic from groundwater has been studied but the inadequate information regarding mechanisms of arsenic removal exists. Ion exchange is not efficient for As(III) removal. Oxidation of this form to As(V) is required before removal. In addition, the production and disposal of spent regenerants and media has also not been studied extensively. Although regenerant production needs to be minimized as much as possible, there is inadequate data regarding the amounts and compositions of the residuals generated by ion exchange processes and methods for disposal. This study involves the determination of the mechanisms of arsenic removal by ion exchange and the characterization of the residuals produced by ion exchange to optimize the conditions for regeneration of the ion exchange media. Factors affecting ion exchange capacity are examined.

2 2.1

MATERIALS AND METHODS Experimental setup

The treatment process consists of an oxidizing filter followed by an ion exchange column. The effect of pH, arsenate concentration and filtration velocity were examined using a simulated Bangladeshi groundwater. The investigation procedure was as follows. The investigation was carried out in the laboratory on a fixed unit incorporating dynamic columns with Filox-R and the anionic-exchange media Purolite A-300 (chloride ion form) serving as oxidizing and ion-exchanging filters respectively. These materials were purchased from Magnor Inc., Boucherville, Quebec, Canada. Purolite A-300 was a Type II, strongly basic gel anion exchange resin. Whole bead counts are a minimum of 92% clear beads with mechanical strengths ranging over 300 grams. Particle size ranged from ⫹16 mesh ⬍5% to 50 mesh ⬍1%. It is unaffected by dilute acids, alkalies, and most solvents. Thus an oxidizing filter was used followed by an anion filtration system. Figure 1A is a schematic diagram of the treatment process used, which consists of the following major elements: Intake: Raw water was pumped from a three-liter flask and flows through the oxidizing filter. A MasterFlex L/S pump with a controller was used. Flow rates were adjusted with the speed control potentiometer as well as an additional flow meter. 239

Oxidizing filter: A plastic syringe was installed, which serves as the oxidizing filter column, to oxidize As(III) to As(V). The oxidizing filter column is a ten milliliter and 1.4 cm diameter plastic column. A MnO2 based material (Filox-R) was used as the oxidizing medium and was regenerated by potassium permanganate solution. FILOX-R™ is the raw, unrefined ore used in the manufacture of FILOX® filtration media. Chemically, FILOX-R (Raw) is naturally occurring ore. Ion exchange system: After passing through the oxidizing filter, water flows into a plastic column filled with the anion-exchange resin Purolite A-300. The ion-exchange column has the same size as the oxidizing filter and the resin bed is 6 mL. Purolite A-300 is a Type II, strongly basic gel anion-exchange resin. 2.2

Regeneration procedure

The oxidizing filter was regenerated with potassium permanganate (133 mg/L) for 30 min at a flow rate of 3 ml/min. This was followed by a slow rinse of 54 ml for 18 min at 3 ml/min and then a fast rinse of 90 ml for 10 min at 9 ml/min. This regeneration is not usually required unless the Filox performance is reduced. The ion-exchange filter (Fig. 1B) was regenerated with 100 ml of potassium chloride solution (2 M) for 33 min at 3 ml/min followed by a slow rinse at the same flow rate for 18 min. Finally a fast rinse was used for 10 min at 9 ml/min. A small amount of caustic (1%) is used in combination with salt during the regeneration in order to enhance the resin operation. This addition gives higher operating capacity. 2.3

Water composition

Simulated water served as the input water. It was made to resemble the real composition of Bangladeshi groundwater (Kinniburgh & Smedley 2001). The following chemicals were used:

• •

H3AsO3 (as specified in the results) FeCl3 ⭈6H2O (0.15 ppm Fe3⫹)

(A)

(B)

Figure 1.

Schematic diagram of the (A) ion exchange setup and (B) regeneration stage.

240

• • • • • •

Al(SO4)3 ⭈14H2O (0.0014 ppm Al3⫹) MnSO4 ⭈H2O(0.001 ppm Mn2⫹) NH4NO3 (0.001 ppm NH⫹ 4) Na2SO4 (0.26 ppm Na⫹) NaF (0.22 ppm F⫺) Distilled water (pH 6.8).

2.4

Experimental procedure

This experiment called for running three factors; namely, input arsenate concentration, pH and filtration velocity, each at two settings, on the ion-exchange process to determine which had the greatest effect on exchange capacity. A (full factorial) 23 design calls for 8 runs was used. Three more runs were set at the center point making 11 runs in total. The design matrix and analysis matrix were prepared before running the experiment; results were recorded in the analysis matrix as every run was completed. Analysis of the results was then done to obtain an equation that describes the dependence of arsenate-exchange capacity on the various factors. The system was stopped as the 10-ppb break through occurred. 2.5

Analytical procedures

The chemical analysis of the samples was performed by a visual method Arsenic detection kit, Hach – 28228-00, (using the Hach method) for the detection of arsenic from 0–500 ppb. This test kit was used for the analysis of most samples. It was purchased from Anachemia Canada Inc. The kit included reaction vessels, chemical reagents (sulfamic acid and powdered zinc) and test strips. Representative samples were also sent to an accredited laboratory for ICP-MS detection. 3 3.1

RESULTS AND DISCUSSION Preliminary test results with As(III)

Concentration (ppb)

Initial tests were performed to determine if As(III) could be removed by the ion-exchange media without oxidation. Using an input concentration of 1600 ppb As(III) into the ion exchange column, it was shown that the concentration of As(III) in the column effluent was approximately the same as the input, indicating negligible removal of this form of arsenic. Subsequently, the oxidizing media was evaluated for its ability to remove As(III). The results are shown in Figure 2. They indicate that the removal of the As(III) was complete until saturation of the column started to occur at a little before 1500 ml. After 2 L of media had passed through the column exhaustion had occurred. Therefore the use of the oxidizing media is a requirement for removal of As(III) as arsenic III needs to be converted to arsenic V prior to treatment. Ion exchange

2000 1500 1000 500 0 0

Figure 2.

1000

2000 Volume (ml)

Sorption of arsenic (III) on oxidizing media.

241

3000

4000

Table 1.

Full factorial design for the sorption experiments, type 23.

Setting

As(V) (ppb)

pH

Linear velocity (cm/min)

Upper Middle Lower

1600 1050 500

8.5 7.5 6.5

1.94 1.30 0.65

Table 2. Determination of arsenate-exchange capacity as a function of initial arsenate concentration, pH and velocity. Operating parameters Initial As(V) Conc. (ppb)

pH

Velocity (cm/min)

Capacity (meq/g)

500 1600 500 1600 500 1600 500 1600

6.5 6.5 8.5 8.5 6.5 6.5 8.5 8.5

0.65 0.65 0.65 0.65 1.94 1.94 1.94 1.94

0.030 0.040 0.010 0.030 0.030 0.030 0.010 0.025

does not remove As(III) because As(III) occurs predominantly uncharged (H3AsO3) in water with a pH value of less than 9.0 (Clifford 1999). The predominant species of As(V) are H2AsO4⫺ and HAsO2⫺ 4 which are negatively charged, and thus are removable by ion exchange. If As(III) is present, this it must be oxidized to As(V) before removal by ion exchange (Clifford 1986). 3.2 Sorption of As(V) Preliminary tests determined the levels and intervals of the factors (Table 1). The concentration range of arsenate 500–1600 ppb is typical of real groundwater in Bangladesh and therefore these levels were chosen for the range of the tests. The velocity range 0.65–1.94 cm/min (giving an empty bed contact time (EBCT) range of 2–6 min) is also typical for most ion-exchange resins. Based on these parameters, 11 runs were conducted. The results are shown in Table 2, it can be seen that the arsenate capacity increases at lower pH values. At the lower pH the ionic form of arsenate H2AsO⫺ 4 would dominate compared to the higher pH where HAsO2⫺ 4 would dominate (Kang & Kawasaki 2000). The resins have been shown to be not sensitive to pH within the range of 6.5 to 9 (USEPA 2003). The difference is more pronounced at the lower arsenate concentration. Increasing the velocity at the lower concentration did not have any significant effects. However, at the higher concentration, the capacity decreased as the velocity increased. The contact time was not sufficient for the higher concentration at the higher velocity. To evaluate the reproducibility of the experiments, the values for the mid-range parameters were evaluated. The capacity of the ion exchange resin for the three repetitions were 0.030, 0.029 and 0.029 meq/g resin, indicating excellent reproducibility. Residual concentrations are shown in Figure 3a, and 3b for the various experiments. The lower arsenate concentrations were achieved at the lowest values for all parameters. In all cases, however, the concentrations were below the current guidelines of 10 ppb. However, if the guidelines 242

Residual As(V) concentration (ppb)

8

a 6 4 pH=6.5 pH=7.5 pH=8.5

2 0 0

1500 500 1000 Input As(V) concentration (ppb)

2000

10 Residual As(V) concentration (ppb)

b 8 6 4 pH=6.5 pH=7.5 pH=8.5

2 0 0

500 1000 1500 Input As(V) concentration (ppb)

2000

Figure 3. Dependence of effluent concentration on pH and initial As(V) concentration. (a) velocity ⫽ 0.65 cm/ min; (b) velocity ⫽ 1.94 cm/min.

As(V) Removal Efficiency (%)

100.0 99.6 High As; Low Velocity

99.2

Low As; Low Velocity High As; High Velocity

98.8

Low As; High Velocity

98.4 98.0 6.5

8.5 pH

Figure 4. Effect of pH on As(V) removal efficiency. High As represents initial concentration of 1600 ppb, low As concentration represents 500 ppb, high velocity represents 1.94 cm/min and low velocity represents 0.65 cm/min.

were lowered to 10 ppb, only lower concentrations of initial arsenate would be able to be treated to achieve acceptable concentrations of arsenate in the effluent. Figure 4 indicates the As(V) removal efficiencies. These rates are higher than the maximum 95% rate according to the USEPA for ion exchange as a Best Available Technology (BAT). 243

Table 3. Sorption and desorption of arsenate from the ion exchange column for lowest concentration of arsenate, pH and velocity.

Bed volumes

As(V) adsorbed (␮g)

As(V) desorbed (␮g)

As(V) removed by regeneration (%)

As(V) in spent solutions (ppm)

Volume of wash (mL)

775

2810

2725

97

11.2

244

3.3

Statistical analysis of the data

Statistical analysis of the data by elimination of insignificant coefficients by Student’s criterion and the adequacy test by Fisher’s criterion was performed and indicated that the pH has the most influence on exchange capacity. The following equation was thus derived from the analysis to relate the ion exchange capacity to the arsenic input concentration, pH and velocity:

(1) where Cas is the initial As(V) concentration and v is the velocity. The values generated by this estimation had a maximum deviation of 2.6% from the experimental one. Regarding the effluent As(V) concentration, the following equation was derived: (2) The maximum deviation between the experimental and estimated values was 7.5% in this case. 3.4

Regeneration of the ion exchange column

The material balance of the sorbed (sorption stage) and desorbed (regeneration stage) arsenate was determined observed (Table 3). The ion-exchange column was regenerated and the spent regenerant analyzed to determine the amount of arsenate the resin sorbed in the sorption process as well as the amount of arsenate that was possible to have desorbed from the exhausted resin in the regeneration stage. This analysis determines the nature of the sorption process, i.e. it gives the answer whether the sorption process is indeed ion exchange and is reversible. An equal amount of arsenate sorbed and desorbed allows make the statement that the sorption process is reversible. In the sorption stage arsenate ions displaced chloride ions but in the regeneration stage chloride ions displaced almost all the arsenate ions and thus recovered the resin as seen in Table 3.

4

CONCLUSIONS

This paper does not attempt to cover the complex of geological, economic and legal problems connected with the remediation of arsenic from groundwater but was limited to the optimization of the ion-exchange technology. There are still many problems that have yet to be solved, such as the effect of sulfate, nitrate and nitrite on sorption capacity and the minimization of the wastes produced during the regeneration of the ion-exchange column. Therefore, attention can be drawn to the following preliminary conclusions of this research.

• •

Feasibility of removing high levels of arsenic concentration at higher than 98% removal was obtained for all conditions evaluated Sorption mechanism is anionic exchange for arsenate ions 244

• • • •

Arsenite cannot be removed without oxidation to arsenate Increasing the pH from 6.5 to 8.5 decreases sorption capability of the resin Polynomial dependencies of arsenate-exchange capacity of different resins on various factors Ion exchange column can be regenerated almost completely as 97% of the arsenic could be removed from the column. Further work will involve optimization of the regeneration processes to reduce waste generation. The same statistical analysis approach will be used.

ACKNOWLEDGEMENTS The authors would like to thank NATEQ for the partial financial support for this project.

REFERENCES Boyle, R.W. & Jonasson, I.R. 1973. The geochemistry of arsenic and its use as an indicator element in geochemical prospecting. J. Geochem. Explor. 2: 251–296. Clifford, D.A. 1986. Synthesis of Dual Mechanism Ion Exchange/Redox Resins and Ion Exchange/Coordinating Resins with Applications to Metal Ion Separations. Discussion of International Symposium on Metals Speciation and Recovery, Chicago, IL, August. Clifford, D.A. 1999. Arsenic Removal Options: Activated Alumina Adsorption, Anion Exchange, and Iron Coagulation-Microfiltration. AWWA Arsenic Technical Work Group Treatment Meeting, Resolve Inc. Washington, D. C., February 18–19. Kang, M. & Kawasaki, M. 2000. Effect of pH on the removal of arsenic and antimony using reverse osmosis membranes. Proceedings of the conference on membranes in drinking and industrial water production, volume 1, pages 489–494. ISBN 0-86689-060-2, October, L’Aquila, Italy: Desalination Publications. Kinniburgh, D.G. & Smedley, P.L. 2001. Groundwater studies of arsenic contamination in Bangladesh. BGS Technical Report WC/00/19. USEPA 2003. Arsenic Treatment Technology Evaluation Handbook for Small Systems. Office of Water (4606M), EPA 816-R-03-014, July. Wang, L, Chen, A. & Fields, K. 2000. Arsenic Removal from Drinking Water by Ion Exchange and Activated Alumina Plants. EPA/ 600/R-00/088, October.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Sorption of arsenic on sorghum biomass: a case study Nazmul Haque, Greg Morrison & Gustavo Perrusquía Water Environment Transport, Chalmers University of Technology, Göteborg, Sweden

Irene Cano-Aguilera, Alberto F. Aguilera-Alvarado & Moisés Gutiérrez-Valtierra Facultad de Química, Universidad de Guanajuato, Guanajuato, Gto., Mexico

ABSTRACT: Large scale, field experiments were conducted for the removal of arsenic from the groundwater of a well located in Guanajuato, Mexico region using non-immobilized sorghum biomass (NISB) as a sorbent, which was found highly efficient to adsorb As in previous laboratory experiments. The columns were run under gravity and pump flow conditions. Removal of arsenic under pump flow was slightly higher than the gravity flow due to the steady-state flow conditions. The maximum arsenic accumulation measured was 3.2 and 3.3 mg of As/g of NISB for gravity and pump flow conditions, respectively. To determine the optimal hydraulic detention time, columns were operated under different flow rates and the maximum sorption occurred at a flow rate of 10 mL/min. Columns of different dimensions were run to obtain the optimal design parameter between surface loading and volumetric loading of the system. The optimal sorption condition can be achieved through the volumetric design of the system.

1

INTRODUCTION

Mining and textile activities can greatly affect the quality of the local groundwater (Bowell et al. 1994, Gray 1997, Rösner 1998). It is more severe in semi-arid lands, where groundwater is the only source of drinking water. High arsenic concentrations in surface and groundwater are recognized to be a problem in mining and textile areas (Gough et al. 1979). It is also recognized that there are potential health consequences for the local population when exposed to drinking water with high arsenic content (Goldman & Darce 1991, Mass 1992, Qvarfort 1992). Uriangato (situated at 20.15°N latitude and 101.18°W longitude) is one of the important centers for textile activities in Guanajuato State of northeastern Mexico (Fig. 1) and, therefore, the city is highly affected by arsenic because these textile industries use a lot of chemicals for their products. One of them might be chromated copper arsenated for the colorization of the cloths. Then, these industries discharge their waste directly to the channel, which passes through the city and links between La Laguna de Cuitzeo and La Laguna de Yuriria as shown in Figure 1 (Virtual Maps & Photos, Mexico 2003). Thus the prevailing hypotheses for the arsenic contamination around this area came from textile industries. A new technology has been evaluated for the removal of arsenic using Sorghum Biomass (SB) as a sorbent (Haque 2003). The laboratory studies using SB both, immobilized and non-immobilized, suggested the following:

• • •

Arsenic can be removed effectively below the guideline value from aqueous solutions using SB. The characterization of arsenic binding to the SB showed that the binding mechanism was pH dependent and the maximum percent of removal of arsenic on SB was at an initial pH of 5.0. As far as the sorption mechanisms are concerned, it is assumed that arsenic is sorbed mainly by an outer surface mechanism. 247

Figure 1.

•

Study area in Uriangato, Guanajuato State in northeastern, Mexico.

SB that was saturated with arsenic showed the remarkable ability for arsenic recovery by treatment with 0.1 M HCl.

This study presents a detailed investigation of arsenic sorption on SB under field conditions. A pilot scale studies were conducted in the field at an arsenic contaminated site in Mexico to elucidate the design factors using NISB as a sorbent in a column to remove arsenic from the aqueous solutions to get the levels consistent for drinking water standards. The columns were run under different flow conditions to determine the maximum efficiency. Optimizations of the columns were studied through the breakthrough curve analysis.

2

EXPERIMENTAL

2.1 Biosorbent The SB from Guanajuato, Mexico region was washed several times with distilled water to remove any particles, or soluble materials because it was waste from the farmers after harvesting their crops. Then, the biomass was dried at 60°C, ground in a bladed mixer to particles less than 0.1 mm in diameter, and sieved through different sieves according to ASTM-AASHO standard methods. These particles were washed with 0.01 M HCl until the supernatant was clear. The biomass was separated from the supernatant then dried in the oven at 100°C for 12 h. This biomass was used for all experiments and it was identified as non-immobilized sorghum biomass (NISB). 2.2 Field experiments The field experiments were conducted at a site called San Jose Curacurio (Fig. 1). The column was packed with NISB. To promote compaction, the column was filled continuously with the suspended biomass in deionized water while the sides of the column were lightly vibrated. The bottom of the 248

Table 1. Running conditions for arsenic sorption to non-immobilized sorghum biomass (NISB) column under gravity and pump flow. Flow condition

Column diameter (cm)

Active bed (cm)

Pore volume (mL)

Amount of NISB (g)

Flow rate (mL/min)

Gravity Pump

5.5 5.5

40 40

535 535

150 150

10 10

Table 2. Running conditions for arsenic sorption to non-immobilized sorghum biomass (NISB) column under different flow rates. Flow rate (mL/min)

Column diameter (cm)

Active bed (cm)

Pore volume (mL)

Amount of NISB (g)

Flow condition

10 20 30

5.5 5.5 5.5

40 40 40

535 535 535

150 150 150

Gravity Gravity Gravity

Table 3. Running conditions for arsenic sorption to non-immobilized sorghum biomass (NISB) column having different sizes. Column diameter (cm)

Flow rate (mL/min)

Active bed (cm)

Pore volume (mL)

Amount of NISB (g)

Flow condition

6.5 5.5 4.5

10 10 10

28 40 60

535 535 535

150 150 150

Gravity Gravity Gravity

column was filled with clean sand (0.25–0.5 mm) to prevent clogging. The column was equipped with a pump and a tank for gravity flow. To estimate the hydraulic detention time of the column, pore volume was divided by the flow rate. The flow rate was observed and adjusted on an hourly basis. All columns had sampling ports that extended into the center of the bed to prevent sampling along the sidewall. Three different sets of experiments were conducted in the field. Firstly, the columns were run under gravity and pump flow conditions to compare the arsenic removal efficiency. Secondly, three columns were run under different flow rates to determine the hydraulic detention time for maximum As removal. Finally, two columns of different sizes but same bed volume were run to test the design parameters surface loading against volumetric loading. The various running conditions for each set of column experiments are shown in Tables 1, 2, and 3. The influent solution was taken directly from the extraction well and was never exposed to the atmosphere prior to passing through the columns. The columns were sampled every half-hour interval at the beginning of the experiments and eventually samples were taken less frequently. The samples were collected from the sample ports in 50 mL glass vials containing few drops of concentrated analytical grade nitric acid (Fisher). A plant influent sample was also collected to determine influent arsenic concentration. On several occasions, samples were collected and analyzed for pH, conductivity, and temperature. Finally, the saturated columns were treated by low concentrated HCl acid to recover the sorbed arsenic from SB. 2.3

Arsenic content determination

Total arsenic content in all experiments was determined by Hydride Generation, Perkin Elmer MHS 15 coupled to Atomic Absorption Spectrometry, Perkin Elmer Analyst 100, (HG-AAS). All 249

the samples were pre-reduced from As(V) to As(III) by adding 10% KI according to the predetermined dilution factor (DF) to prevent the interferences between the two oxidation states of arsenic. Samples were stored at least 30 min in darkness and were analyzed within 3 h of sampling. Each time a 1 mL sample was pipetted into a 50 mL volumetric flask where 10 mL 1.5% of HCl was added for dilution. Subsequent inline hydride generation with argon and sodium borohydride reductants (3% NaBH4 in 1% NaOH solution) and HG-AAS detection allowed detection of surely total arsenic species down to 10 ␮g/L at a wavelength of 193.7 nm. The signals for each sample were read to provide a mean and relative standard deviation. Calibrations curves were performed for the removal range of analysis (10 to 5000 ␮g/L) and a correlation coefficient for the calibration curve of 0.986 or greater was obtained. The instrument response (50 ␮L of the 1000 mg/L arsenic stock solution give an absorbance of approximately 0.2) (Perkin Elmer 2000) was periodically checked with known arsenic standards. 2.4 Error analyses The adsorption experiments were carried out in triplicate in order to evaluate the experimental reproducibility of the experimental results. The confidence of data generated in the present investigations has been analyzed by standard statistical methods to determine the mean values and confidence intervals. Each data set was calculated at 95% confidence level (P ⬍ 0.05) to determine the error margin (William & Ricahard 1973). The correlation coefficient was computed as required to confirm the linear range for a minimum of 12 data points.

3 3.1

RESULTS AND DISCUSSION Field experiments

The influent arsenic concentrations for the entire field study are shown in Figure 2. The variability range from 471 to 541 ␮g/L with an average concentration of 506 ␮g/L. An influent sample was taken during each sampling event. The first phase of the column experiments was conducted separately to compare the capacity under gravity and pump flow condition. Effluent arsenic concentrations for both flow conditions are shown in Figure 3. Arsenic removal below the guideline value (10 ␮g/L), recommended by World Health Organization (WHO 2001), was observed for up to 475 and 500 pore volumes for the gravity and pump flow conditions, respectively. Breakthrough (⬎65 ␮g/L and ⬎60 ␮g/L for gravity and pump flow, respectively) occurred after 333 and 400 h for gravity and pump flow, respectively. Then the columns slowly decreased in their ability to remove arsenic from the solution and the column under gravity flow saturated after 1150 pore volumes and the column under pump flow after 1200 pore volumes. The maximum sorption capacity and the supply of fresh water (below the guide line value provided by WHO) for the column under gravity flow were 3.2 mg of As/g of NISB and 254 L and for the column under pump flow were 3.3 mg of As/g of NISB and 267 L. The slight difference of results between the two flow conditions is attributed due to the variability in flow. A steady state flow was achieved under pump flow condition which might be obtained under gravity flow condition if the water level can be kept constant in the tank. It was necessary to measure the pH, redox potential, and the temperature of the influent solutions. Because all these parameters are important for sorption of arsenic to SB. From the pH profile experiments (Haque 2003), the results showed a strong influence of pH and the maximum removal of arsenic was observed at an initial pH range of 4.0–6.0. During the field experiments, the initial pH ranged from 4.6 to 6.4 which were favorable for the sorption of arsenic to SB. However, it might not be same for other conditions where the solutions pH changes a lot. The second set of the column experiments were carried out to determine the hydraulic detention time. The breakthrough curves for different flow rates are shown in Figure 4. It can be easily seen from the breakthrough curves that arsenic can be sorbed more with less flow rate. From the previous 250

600 500

µg/L

400 300 200 100 0 0

200

400

600

800

1000

1200

Pore Volumes

Figure 2.

Influent arsenic concentrations of the natural source.

1 Pump Flow

C/Co

0,8

Gravity Flow

0,6 0,4 0,2 0 0

200

400

600 800 Pore Volumes

1000

1200

Figure 3. Breakthrough curves under pump and gravity flow conditions. The flow rate of the incoming As solution into the column was 10 mL/min.

1,0 10 ml/min 0,8

20 ml/min

C/Co

30 ml/min 0,6 0,4 0,2 0,0 0

Figure 4.

200

400

600 800 Pore Volumes

1000

1200

Breakthrough curves under different flow rates. The columns were run under gravity flow.

251

1 Column Dia. 6.5 cm Column Dia. 5.5 cm

0,8

Column Dia. 4.5 cm C/Co

0,6 0,4 0,2 0 0

200

400

600

800

1000

1200

Pore Volumes

Figure 5. Breakthrough curves among the different sizes of column but same bed volumes. The flow rate of the incoming As solution was 10 mL/min under gravity flow.

sorption isotherm analysis (Haque 2003), it was noticed that the arsenic binding was rapid within 1 hour. The rapid binding of the metal ions by SB could indicate the metals are being adsorbed onto the surface of the biomass, instead of absorbed. It has been found that the adsorption processes in most of the biomasses are within 5 minutes and only the surface areas of the sorbent are highly active during this time period (Vaishya and Prasad 1991). High flow rate might break the binding surface between arsenic and SB and, therefore, the sorption was less. The final set of column experiments were carried out to determine the design parameter surface or volumetric loading. The size of the columns was different but the total volume of the sorbent was same. Figure 5 shows the breakthrough curve among the different sizes of column but same bed volumes. The breakthrough curves were at 200, 450, and 650 pore volumes for the column diameter of 6.5, 5.5, and 4.5 cm, respectively. The sorption capacity of the column increased with the lower diameter i.e. the higher length of the column. So, the efficiency of the column can be increased by volumetric design but the ratio of column length to column diameter should be ⬃5 to avoid plugging of the column and other undesirable effects. From the batch experiment (Haque 2003), it has been found that the recovery of arsenic from the saturated biomass is achievable. Thus, it was intended to recover sorbed arsenic from the saturated column by the addition 0.1 M HCl. Approximately 95% of the arsenic bound was recovered from the column as the acid was passed through it. The little arsenic remaining may be responsible for the slight decrease on arsenic binding capacity, when the column will be used again.

4

CONCLUSIONS

This research clearly demonstrated the effectiveness of SB to remove arsenic under natural conditions. More specific conclusion can be drawn from the results of each specific experimental case:

• • • • •

Due to the steady-state flow condition, arsenic removal under pump flow was slightly higher than the gravity flow. The maximum arsenic accumulation measured was 3.2 and 3.3 mg of As/g of NISB for gravity and pump flow conditions, respectively. The sorption capacity of the column can be optimized through the lower flow rate (10 mL/min) without alteration on the efficiency of the column. The system can be designed on a volumetric loading basis. The saturated column can be reused after recovery of bound arsenic by using 0.1 M HCl. 252

Finally, this case study showed that the groundwater of Guanajuato, Mexico region could be treated for drinking purposes by using SB. The adsorbents are not only inexpensive but also available. This innovative technology provides a reusable material (SB) which is not biodegradable, but environmentally friendly.

ACKNOWLEDGEMENTS The Swedish Foundation for International Cooperation in Research and Higher Education (STINT) are greatly acknowledged for their financial support.

REFERENCES Bowell, R.J., Morley, N.H. & Din, V.K. 1994. Arsenic speciation in soil pore waters from the Ashanti Mine, Ghana. Appl Geochem 9: 15–22. Fotos y Mapas Virtuales Departmento, Mazatlan, Mexico 2003. Ave. Cruz Lizarraga, No. 712–3, Col. Palos, Prietos, Sinaloa, Mexico 82100. Goldman, M. & Darce, J.C. 1991. Inorganic arsenic compounds: are they carcinogenic, mutagenic, teratogenic? Environ Geochem Health 13: 179–191. Gough, L.P., Shacklette, H.T. & Case, A.A. 1979. Element concentrations toxic to plants, animals, and man. USGS Bull 1466. Gray, N.F. 1997. Environmental impact and remediation of acid mine drainage: a management problem. Environ Geol 30: 62–71. Haque, N. 2003. Sorption of arsenic on Sorghum Biomass and Iron Fillings. Master Thesis. Department of Water Environment Transport, Chalmers University of Technology, Göteborg, Sweden. Mass, M.J. 1992. Human carcinogenesis by arsenic. Environ Geochem Health 14: 44–54. Perkin Elmer 2000. User’s Guide for Mercury Hydride System, USA. Qvarfort, U. 1992. The high occurrence of arsenic in Macadam products from an iron mine in central Sweden: significance for environmental contamination. Environ Geochem Health 14: 87–90. Rösner, U. 1998. Effects of historical mining activities on surface water and groundwater: an example from northwest Arizona. Environ Geol 33: 224–230. Vaishya, R.C. & Prasad, S.C. 1991. Adsorption of copper (II) on sawdust, Indian J. Environmental Protection. 11(4): 284–289. WHO 2001. Arsenic compounds: In Environmental Health Criteria 224, 2nd edition, World Health Organization, Geneva. William, M. & Ricahard, L.S. 1973. Mathematical Statistics with Applications. Duxbury Press, Massachusetts: pp. 274–290.

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Removal and recovery of arsenic from aqueous solutions by sorghum biomass Z.I. González-Acevedo, I. Cano-Aguilera & A.F. Aguilera-Alvarado Facultad de Química, Universidad de Guanajuato, Guanajuato, Gto., México

ABSTRACT: This work focuses on the development of bio-filters for the removal and recovery of arsenic (As) from aqueous solutions, based on the natural properties of sorghum biomass to adsorb metals. The As adsorption capacity at pH 4.5 and 60 min of contact time was 8.43 mg/g and 6.99 mg/g for free and immobilized biomass, respectively. After HCl treatment, As was recovered in 91% and 95% from free and immobilized As saturated biomass, respectively. In flow conditions, the adsorption’s capacity of the immobilized biomass packed in a column was 143 mg/cm3 when the input rate of the arsenic solution was 3 mL/min. The designed column size to remove As from a well with high As content, taking in consideration several factors, was 16 cm diameter and 80 cm height. Using two columns in series and one of reserve, a previous cost analysis showed that the water cost after this treatment was 0.002 and 0.011 US$/L for free and immobilized biomass, respectively.

1

INTRODUCTION

Freshwater is a renewable but limited resource. The growing demand of clean and safe water from the population, the industry and the agriculture has been taking conscience that the practices of the past, which supposed an infinite source of water at low price, cannot longer continue. The human activity has affected the quantity and quality of several water bodies in the world. The growing demand had lead many countries to test the best way to make a well use and to preserve the water reservoirs to future uses in the context of a sustainable development (Henry 1996). Groundwater reservoirs are a principal source of fresh water. The world grounds aquifers layers conform 90% of the total fresh water, that could be use to human consume. The groundwater in Mexico is an important national resource, because more than 95% of the rural population uses this kind of source to cover their necessities (CEAS 1998). In Mexico, arsenic has been founded in groundwater above of the permissible limits of the Mexican Official Norms (NOM-127-SSA1 2002) which is 40 ␮g/L. This limit should decrease each year 0.05 ␮g/L until to reach 10 ␮g/L, limit of the World Health Organization (WHO 2001). In Mexico the more affected states are: Hidalgo, Morelos, Coahuila, Durango, Guanajuato, Zacatecas and Chihuahua (Armienta 1997). The origin of As could be natural or anthropogenic (Petkova et al. 1998). In the State of Guanajuato the highest concentration of arsenic detected was 300 ␮g/L, and its presence is affecting the health of population. The intoxication is more evident, especially in certain areas (Cano-Aguilera 1999). Long-term exposure to arsenic in drinking water is causally related to increased risks of cancer in the skin, lungs, bladder and kidney, as well as other skin changes such as hyperkeratosis and pigmentation changes (WHO 2002). There exist several conventional methods to remove metals in solution (chemical precipitation, chemical oxidation or reduction, filtration, electrochemical treatment, application of the membranes technology and recovery by evaporation) these methods could be ineffective or economically prohibitive, specially when the metals are dissolved in big volumes of relative low concentration solutions (around 1–100 mg/L) (Volesky 1987). 255

The use of biological materials with the capacity to remove and to recover metals has been proposed as an alternative process, because of its low price (Drake & Rayson 1996), the minimization of secondary problems that the sludge could cause, and its high efficiency in diluted effluents (Schneegurt et al. 2001). These processes apply active, inactive or modified natural materials (biomaterials, biomass), which have the capacity to adsorb diluted metallic ions. The adsorption processes by biomaterials are based on the natural and strong affinity of their cellular compounds with the metallic ions (Wang 2002). Because of these facts, the propose of this work is the research and development of a new clean efficient and low cost technology, based on the bio adsorption processes to remove and recover As from aqueous solutions, focus in the use of industrial sorghum waste as bio adsorbent.

2

METHODOLOGY

Several parameters were evaluated, to find the conditions where the sorghum biomass had a better As uptake and which were useful to make the scale up. Some of these parameters were evaluated in batch and some others in column experiments. 2.1

Preparation of free and immobilized biomass

To prepare the free biomass, the industrial waste of sorghum (biomass) was carefully washed with a solution of HCl (0.01 M), then it was dried at 120°C during 24 h and then sieved into a fine powder (mesh 100 ⫽ 0.1 mm of particle diameter). In terms of operability of the bio filters it is recommended to immobilize the biomass. The procedure followed was: in 75 mL of H2SO4 (5%) was added Na2SiO3 (6%) until to reach pH 2, then were added 2 g of free biomass in continuously agitation, after that, was added more solution of Na2SiO3 (6%) until to reach pH 7 when the polymerization process started and the incorporation of the biomass into this polymer began, after that the immobilized biomass was washed with enough water to eliminate sulfates (that was verified with drops of 5% solution BaCl2. Then the immobilized biomass was dried at 120°C during 48 h and sieved (mesh 60 ⫽ 0.25 mm of particle diameter). 2.2 Batch conditions When the biomass was free and immobilized, the parameters that were evaluated for both of them were: pH, contact time (between the As solution and the biomass), saturation capacity of the biomass and desorption process. For these experiments the concentration of biomass solution was 5 mg/mL, each time the pH was adjusted, the tubes centrifuged and the supernatants were separated from the rest of the biomass. Then the arsenic solution or HCl solution (for desorption process) was added to the biomass in the tubes and they were agitated during the contact time, after that, the tubes were centrifuged again and the supernatant separated to determine the arsenic concentration. All the analysis of total As in solution were determined by Flame Atomic Absorption Spectroscopy (FAAS) in an Atomic Absorption Spectrometer Perkin Elmer Model Analyst 100. All the experiments were performed 3 times and a control without As was introduced. The solutions were prepared from a 1000 ppm As Perkin Elmer standard and deionized water. 2.3 Continuous conditions To make the experiments at continuous conditions, (showed in Fig. 1), glass columns of 9.5 cm height and 1 cm diameter were packed with immobilized biomass (3), the rate flow was 3 mL/min (it was controlled with a peristaltic pump (2)), before each experiment; the column was stabilized with HCl pH 4.5 (let it running during 5 min). After the stabilization, the As solution (6.3 ppm) at 256

Figure 1. As removal by sorghum biomass in continuous conditions. (1) As solution or HCl solution pH 4.5, (2) peristaltic pump, (3) bio column, (4) automatic fraction collector.

pH 4.5 was run through the column. The volume bed was collected with an automatic BIO-RAD 2110 fraction collector, the fractions were collected with 80 bed volumes each round. Fractions of 20 tubes were taken (indiscriminate) to analyze them and to obtain the loading curve of the adsorption system arsenic-sorghum. With this curve, the results from batch conditions and the parameters from a particular case, the scale up of the bio filters was made as well as the preliminary cost analysis. This experiment was made for a single column and for two columns in series. For the columns in series experiment, 2 columns (previous stabilized) were set up; one over the other and when one of the columns was saturated, then the column one was replaced with the column two and the column two with a new column (previous stabilized), in order to have a continuous process.

3 3.1

RESULTS AND DISCUSSION Batch conditions

First the pH at which the sorghum biomass adsorbed the highest quantity of As in solution was evaluated. At pH 4.5, between 70 and 80% of the total arsenic from the solution removed. For the free as well as the immobilized biomass, the pH value considered was 4.5 (Fig. 2). The contact time at which the best arsenic adsorption occurs when the biomass is free and immobilized and at the predetermined pH of 4.5, as the Figure 3 shows, is 60 min of contact time between the biomass and the arsenic solution, adsorbing a little bit more than 80% of the total amount of As in the solution and the immobilized biomass around 70%. After this time the arsenic adsorption is constant during large periods of time. The saturation capacity of the sorghum biomass, in other words, the maximum quantity of As ions that the biomass could uptake, was evaluated in both; the free sorghum biomass and the immobilized one, at the pH and contact time previously determined, resulting 8.43 mg of As/g of dry biomass when the biomass is free and 6.99 mg/g for the immobilized one (Fig. 4). For As recovery, the sorghum biomass As saturated was treated with a HCl 0.1 M solution. Figure 5 shows that the desorption process for the free and immobilized biomass during the first additions was able to recover the major quantity of As, and this quantity decreases when the recovery process was occurring. The percent of As recovered for free biomass was 91% and for immobilized one was 95%. 257

100 % As removed

80 60 40 20 0 0

1

2

3

4

5

6

7

pH

% As adsorpted

Figure 2. Adsorption of As by the sorghum free (•) and immobilized (䉱) biomass as a function of the pH. The concentration of As solution was 6.3 ppm and 60 min of contact time in continuous agitation. The biomass concentration was 5 mg/mL. 90 80 70 60 50 40 30 20 10 0 0

20

40

60

80

100

Time (min)

Figure 3. Adsorption of As to free (•) and immobilized (䉱) sorghum biomass as a function of the contact time. The concentration of the arsenic solution was 6.3 ppm in continuous agitation and pH 4.5. The concentration of biomass was 5 mg/mL. 60 % As removed

50 40 30 20 10 0

0

1

2

3 4 Addition Number

5

6

7

Figure 4. As saturation capacity for the free (•) and immobilized (䉱) sorghum biomass. The concentration of the As solution used in each addition was 6.3 ppm, pH 4.5 and 60 min contact time. The concentration of the biomass was 5 mg/mL.

3.2

Continuous conditions

For flow conditions, the As adsorption capacity of an experimental bio-column (9.5 cm height and 1 cm diameter) packed with immobilized biomass and stabilized at pH 4.5 was evaluated. The saturation of the column was reached after 400 fractions (bed volumes), equal to 13.3 h and 2.4 L of arsenic solution passed through it (Fig. 6). 258

80

% As recovered

70 60 50 40 30 20 10 0 0

1

2

3 4 Addition Number

5

6

7

Figure 5. As recovery from the As saturated sorghum free (•) and immobilized biomass (䉱) with acid treatment. The concentration of HCl was 0.1 M and 5 min of contact time. 7 ppm of As in the effluent

6 5 4 3 2 1 0

0

100

200 300 Bed Volume

400

500

Figure 6. Adsorption of As on an experimental bio column. The As solution concentration was 6.3 ppm and pH 4.5. The solution input rate was 3 mL/min and bed volume or collected fractions were 6.3 mL.

The system in series shows that in the first fractions, the adsorption process does not reach a maximum value; but rather maintains an approximated value of 50%, after 400 fractions. This agrees with the result obtained for a single column, the expectation was to find a saturation after this point (400 fractions), but the behavior was constant and it was decided to collect 80 more fractions, looking then a decreasing of the adsorption and according to this, the column 1 where replaced for the column 2 and a third column previously stabilized was placed in the position of the column 2. At this point, there is an increasing adsorption that is represented for a decreasing of the As concentration in the effluent (Fig. 7).

4

APPLICATION

With the previous results, was made the determination to the bed dimensions for a column that could remove As from natural water with specific characteristics for a case study (a well from the locality of San José Curacurio, in the State of Guanajuato). This well supplies water for 240 families and the water demand is 36 L/day per family, and the total water to be treated is 360 L/day. Considering that the system work at 60% of efficiency, because of possible changes of pH, 259

10 ppm of As in the effluent

9 8 7

New column

6 5 4 3 2 1 0 0

100

200

300

400 500 600 Bed Volume

700

800

900

1000

Figure 7. Saturation of the series of 2 columns system with As. The concentration of the As solution was 6.3 ppm, pH 4.5 the input rate was 3 mL/min and the bed volume or the volume of the collected fractions was 6.3 mL. The saturated column was replaced after 480 fractions.

temperature, As concentration, etc., a column of 16 cm of diameter and 70 cm height with a bed volume of 13,988 cm3 and the exhausting time is 22 months was designed. The preliminary cost analysis showed that for a system in series (3 columns of polycarbonate) the initial investment is 22,500 US$ when the biomass is immobilized and 3000 US$ when the biomass is free. The maintenance cost per year is 5300 US$ when the biomass is immobilized and 800 US$ when is free. In case of recover the As, the process would cost 700 US$/year and the recovery cost for As selling would be 113 US$/year. Therefore, the minimum cost of the treated water is 0.011 US$/L when the biomass is immobilized and 0.002 US$/L when it is free.

5

CONCLUSIONS

The biomass of industrial waste of “sorghum” was able to adsorb As from aqueous solutions with a capacity of 8.43 mg/g and 6.99 mg/g for the free and immobilized biomass respectively, at pH 4.5 and 60 min of contact time. After the immobilization with sodium silicate, the As biomass uptake decreased 10% approximately. The sorghum biomass saturated with As and lately treated with HCl 0.1 M solution was able to desorb As in 91% and 95% from the free and immobilized biomass, respectively. A column (9.5 ⫻ 1 cm) packed with immobilized sorghum biomass had the capacity to adsorpt 143 mg of As per cm3 of immobilized bed, when the water flow rate is 3 mL/min. From the saturation results of the single column and the water quality of a particular well, the column size was estimated for several operation conditions. The best recommended size to this process and carried out with 60% efficiency, was 16 cm diameter and 80 cm height. The proposal based in the case study, was to use two columns in series and one of reserve to treat the water coming from the well. A preliminary cost analysis for As removal and recovery processes from aqueous solutions indicates a cost of 20 ¢ (US$) per liter of treated water, in function of the volume to be treated.

REFERENCES Armienta, M.A., Rodríguez, R., Aguayo, A., Ceniceros, N., Villaseñor, G. & Cruz, O. 1997. Arsenic Contamination of Groundwater at Zimapán México. Hydrogeology Journal 5(2): 39–46. Cano-Aguilera, I. & Rodríguez, E. 1999. Estudio de la Presencia de Arsénico, Selenio y Plomo en Agua de Consumo de los Pozos del Estado de Guanajuato. Reporte Técnico, Universidad de Guanajuato, México.

260

CEAS (Comisión estatal de agua y saneamiento de Guanajuato) 1998. Sinopsis del estudio hidrogeológico y modelo matemático del acuífero del Valle de Pénjamo-Abasolo, Gto. Drake L.L. & Rayson D.G. 1996. Plant-Derived Materials for Metal Ion – Selective Binding and Pre concentration, Analytical Chemistry News and Features, January 1. Henry J.G. & Heinke G.W. 1996. Ingeniería Ambiental. Segunda Edición, Prentice Hall. Petkova S.V., Rivera H.M de L., Piña S.M., Avilés F.M. & Pérez C.S. 1998. Evaluación de Diversos Minerales para la Remoción de Arsénico en Agua para el Consumo Humano. IMTA. Ingeniería y Ciencias Ambientales 34. NOM-127-SSA1 2002. Diario Oficial de México. Schneegurt, A.M., Jain, J.C., Menicucci, J.A. Jr., Brown, S.A., Kemmer, M.K., Garofalo, F.D., Quallick, R.M., Neal, R.C. & Kulpa, F.C. Jr. 2001. Biomass Byproducts for the Remediation of Wastewaters Contaminated with Toxic Metals, Environmental Science and Technology 35(18): 3786–3791. Volesky, B. 1987. Biosorbents for Metal Recovery. Volume 5, Elsevier Publications, Cambridge, Tibtech. Wang, J. 2002. Biosorption of copper (II) by chemically modified biomass of Saccharomyces Cervisiae, Process Biochemistry 37(8):847–850. WHO (World Health Organization) 2001. Arsenic in Drinking Water; Informative note. WHO (World Health Organization) 2002. Health Criteria No. 224; Arsenic.

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Optimisation of iron removal units to include arsenic removal A.K. Sharma, J.C. Tjell & H. Mosbæk Environment & Resources, Technical University of Denmark (DTU), Lyngby, Denmark

ABSTRACT: Elevated concentrations of arsenic (As) along with iron (Fe) often occur in the ground waters in flood plains of West Bengal (India) and Bangladesh. In practice As can be removed when Fe is added as coagulant after oxidation of As(III) to As(V) and adsorbing As(V) onto Fe hydroxides. In Bangladesh iron removal units (IRU) were installed to remove naturally occurring Fe to supply Fe-free water to villagers. Hence investigations were carried out to examine the efficiency of IRU for As removal and to study whether similar As removals can be obtained by simple sedimentation in buckets compared to the IRU. The results show that IRUs have removal rates for As comparable to simple sedimentation. It was observed that Fe/As molar ratios above 80 were necessary to achieve residual As concentrations below 50 ␮g/L. Based on the extensive data available on As and naturally occurring Fe concentrations in wells, removal of As may be attained at a quarter of the tube wells exceeding an As concentration of 50 ␮g/L. The results further showed that due to improper construction of the IRUs, untreated water was mixed with treated water and hence resulted in poorer Fe and As removals. It is concluded that removal of As by naturally occurring Fe in IRUs in the areas with elevated Fe concentrations is functioning and can be employed as an As removal method, but further research is needed to improve the construction of the IRUs with a possibility of introducing a simple As(III) oxidation unit and to improve the Fe removal capacity of the IRU and thereby increase the As removal efficiency.

1

INTRODUCTION

Presence of iron (Fe) in the ground waters of Bangladesh is known for decades and efforts have been carried out by many international funding agencies and local authorities to reduce the content of Fe in the water for domestic usage. The problems of using Fe contaminated groundwater for domestic purposes are laundry staining, bad taste, un-aesthetic colour and smell, corrosion of pipes and pumps, deposits on utensils, discoloration of rice etc. According to the Bangladesh standard 1 mg/L of Fe is allowed in the drinking water in rural areas and 0.3 mg/L of Fe in towns where piped water supply systems are available. Since only Fe2⫹ is soluble in the pumped water from tube wells one of the easiest ways is oxidation through exposure to air and subsequent removal of the formed Fe-(hydro)-oxides by sedimentation and filtration. Based on this principle several removal methods at domestic and community levels have been developed. Groundwater is gradually being introduced as a safe drinking water source to the population of Bangladesh since about 3 decades as an alternative to microbiologically contaminated surface water. However, the detection of arsenic (As) in the ground waters in 1993 created a huge need for supply of As free drinking water to a population above 40 million, that are believed to be exposed to As concentrations above 50 ␮g/L, which is the present maximum contamination level (MCL) for drinking water in Bangladesh (Smedley & Kinniburgh 2002). As is classified as a human carcinogen. WHO lowered the recommended MCL of 50 ␮g/L to a provisional guideline concentration of 10 ␮g/L in 1993 (WHO 1996). Available As removal methods can be classified into four categories: coagulation/co-precipitation, ion exchange, adsorption, and membrane separation, among which coagulation/co-precipitation is considered to be a simple and low cost removal method that can easily be adopted in developing 263

Table 1. Tube wells with arsenic concentration above 50 ␮g/L exceeding a given Fe/As ratio (M/M), (data from BGS 1998). Fe/As ratio (M/M)

% Tube wells exceeding

10 50 100

81 35 16

countries like Bangladesh. The only drawback is production of sludge, which has to be disposed of properly. The common coagulants used for the removal of As are locally available salts of Fe and aluminium. However since most of the ground waters in Bangladesh contain high Fe concentrations, the presence of naturally occurring Fe may be taken advantage of. Most reported As concentrations in Bangladesh are below 0.5 mg/L while Fe concentrations in excess of 5 mg/L are common for many regions of the country (Hossain & Ali 1996). Nickson et al. (1998) have demonstrated a positive correlation between Fe (⬍26 mg/L) and As (⬍0.26 mg/L) in groundwaters in Bangladesh. According to Ahmed et al. (2004) total Fe exhibits a relatively low correlation with total As (r2 ⫽ 0.42) and this can be explained by precipitation of siderite and vivianite. Table 1 shows the proportion of tube wells exceeding increasing Fe/As ratios. The values are calculated based on the extensive data collected by BGS (BGS 1998). Removals of As with Fe depend on many factors like the ratio of As(III)/As(V), pH, and presence of competing anions (Jain & Loeppert 2000, Hering et al. 1996, Su & Puls 2001, Meng et al. 2000, Pierce & Moore 1982, Raven et al. 1998, Gao & Mucci 2001, Manning & Goldberg 1996). The available literature survey shows that As(V) can be easily removed compared to As(III), and according to the literature survey the ground water of Bangladesh contain both As(III) and As(V) resulting in the need of pre-oxidation of water for effective removal of As. Oxidation of As(III) can be achieved by oxidants like free chlorine, hypochlorite, ozone, permanganate, solid manganese dioxide, and hydrogen peroxide/Fe2⫹ (Borho & Widerer 1996). Fe in presence of oxygen and/or complexing agents like citric acid in presence of sunlight (Hug et al. 2001) also results in oxidation of As(III) to As(V). Best removals of As(V) were achieved at pH between 6–8. The presence of anions like phosphate, silicate, sulphate and bicarbonate decreases As removal resulting in higher coagulant dosages. According to Meng et al. (2000), the Fe/As ratio required for removal of As(V) below 50 ␮g/L increases from 12 in the absence of phosphate and silicate to above 40 when these ions are present at respective concentrations of 1.9 mg/L and 18 mg/L (average ion concentrations in Bangladesh tube well water). In many areas with elevated Fe concentrations, Iron Removal Units (IRU) are installed to supply Fe free water from wells to villagers. Since Fe is added as coagulant for removal of As and that Fe can oxidize As in presence of oxygen these removal units have a great potential for As removal. Limited literature is available on the functioning of these IRUs for As removal. Therefore the present work investigates the functioning of the IRUs as As removal units and suggest which measures are to be taken to improve the effectiveness of IRUs for removal of As below the present Bangladesh MCL for drinking water of 50 ␮g/L.

2

MATERIALS AND METHODS

To supply Fe free water to the villagers of Bangladesh DANIDA (Danish International Development Agency) has on a trial basis supported construction of IRU in two villages: Lakshmipur and Choumohani. Since groundwater in these areas also have high As concentrations they were chosen for the field investigations and 10 IRU were chosen for the present study with the number of people using each IRU varying between 7–300. Two of the IRU were sampled twice to see whether there would be differences in removals over time. 264

Inlet

Aeration

Treated water

Outlet

Filtration

Sedimentation

Figure 1.

Iron removal unit (not to scale).

The main principle of an IRU (Fig. 1) is aeration, sedimentation and filtration of the pumped raw groundwater. The study was on the simultaneous As removal at different stages of the IRU where water samples were collected from raw water, after aeration, and treated water. Comparative experiments were also conducted to study the removal of As by plain sedimentation of the iron-hydroxides in 1 L plastic bottles to investigate the efficiency of IRU compared to plain sedimentation in buckets. The mode of plain sedimentation experiments was to let water containing elevated concentrations of As and Fe to stand for a certain period and collect the samples at the end of the sedimentation period. Similar sedimentation experiments were also conducted in West Bengal to see the effect of time of sedimentation on As removal by allowing 10 L of As contaminated water in open plastic buckets to stand for 24 hours. Samples were collected after 3, 8 and 24 hours of sedimentation.

2.1

Collection, preservation and analysis of samples

Samples from the IRU were collected in 1L plastic bottles pre-washed with the respective samples and preserved by adding 14 M HNO3 (1 mL/L). For the analysis of As(III) and As(V) the Silver Diethyl Dithiocarbamate method (APHA 1995a) was used and the absorbance was measured on filter photometer (PF-10) at 520 nm. The standard deviations for As determination was less than 10% for As concentrations above 50 ␮g/L. The samples from West Bengal were measured for total As using a slightly modified hydride generation-AAS method (APHA 1995b). The thioglycolic acid method (HMSO 1972) was used for the analysis of total Fe and the absorbances were measured on a spectrophotometer at 535 nm. Phosphate was measured using the Automated Ascorbic Acid Method (APHA 1995c). Silicate and total Fe were measured using Flame-AAS (APHA 1995d). 265

3

RESULTS AND DISCUSSION

The concentration of Fe in the raw water at the selected IRU ranged between 6.7–17.7 mg/L with an average of 12 mg/L, whereas the concentration of As was in the range of 0.18–0.5 mg/L with an As(III)/Total As ratio in the range of 0–0.56. Fe/As ratios (M/M) were then within 20.6–127.3. The phosphate and silicate concentrations were not measured but according to BGS the concentrations in these areas are in the range of ⬍0.1–7.8 and 9–40 mg/L respectively. 3.1

Removal of Fe in the IRU and after plain sedimentation

The residual Fe concentrations in the treated water with IRU and plain sedimentation are presented in Figure 2, and the results indicate that higher Fe removals were obtained in the plain sedimentation experiments compared to in the IRU. The reason behind the poorer Fe removals by the IRU could be construction of the IRU, which results in mixing of untreated water with treated water. The field survey of the IRU showed that between 0–50% of the raw water was directly led to the treated water tank at most of the IRU. The results from the plain sedimentation experiments and the field observations indicate that by proper construction of the IRU where the raw water is treated first before entering into the treated water, better Fe removal can be achieved. 3.2

Removal of As in the IRU and after plain sedimentation

Figure 3 shows the residual As concentration in the IRU and plain sedimentation as a function of the initial Fe/As (M/M). The obtained results indicate that the removal of As with IRU is comparable to the removal of plain sedimentation with few exceptions. Differences in observed performance may stem from that samples from plain sedimentation experiments were not filtered and hence with a possibility of transfer of iron hydroxide particles loaded with As during sampling. Figure 3 shows that it is possible to achieve a residual As concentration below 50 ␮g/L when the Fe/As ratio is above 80. Based on the extensive data from BGS, around one quarter of the tube wells having As concentrations above 50 ␮g/L exceeds an Fe/As ratio of 80. Meng et al. (2001) achieved the 50 ␮g/L at Fe/As ratios below 40 by oxidising As(III) with NaOCl. Mamtaz & Bache (2000) reported that an Fe/As ratio of 40 and a settling time of 3 days was necessary to achieve a

3.50

Final Iron (mg/L)

3.00

Plain Sedimentation IRU

2.50 2.00 1.50 1.00 0.50 0.00 0.00

5.00

10.00

15.00

20.00

Initial Iron (mg/L)

Figure 2. Relationship between concentrations of initial Fe in raw water, and water treated in the IRU or in plain sedimentation.

266

residual As concentration below 50 ␮g/L when As is present as As(III) and in the absence of competing ions. The main reasons behind the higher Fe/As needed here are the slow oxidation of As(III) and the presence of competing anions. In this study As is partly present as As(III) and there are high concentrations of interfering ions in the water. To investigate whether the residual As in water is due to incomplete oxidation of As(III), the samples were also measured for As(III). Figure 4 shows the residual As(III) concentration after aeration and treatment in the IRU, or after plain sedimentation, as a function of the initial As(III) in raw water. The obtained results indicate that about 85% of the initial As(III) was present after aeration with a correlation coefficient of 0.81. For treated water in the IRU and by plain sedimentation the concentration of As(III) present was below 10 ␮g/L except at two IRUs, but in all cases the residual As(III) present was below

Residual Arsenic concentration (mg/L)

0.35 IRU Plain sedimentation

0.30 0.25 0.20 0.15 0.10 0.05 0.00 0

20

40

60 80 Fe/As (M/M)

100

120

140

Figure 3. Residual As concentration in treated water by IRU or in plain sedimentation as a function of the initial Fe/As ratio.

As(III) concentration in the processed water (mg/L)

0.25

0.20

0.15 y = 0.86x R2 = 0.81 0.10

0.05

0.00 0.00

Figure 4.

After aeration in IRU After treatment in IRU After Plain sedimentation

0.05

0.10 0.15 0.20 0.25 As(III) concentration in the raw water (mg/L)

Oxidation of As(III) at different stages in the IRU and in plain sedimentation.

267

0.30

50 ␮g/L. These results indicate that the residual As present in the treated water is not due to incomplete oxidation of As(III) and thus not the main reason for the poor As removals in the IRU. Other reasons for inefficient removal of As could be insufficient treatment time, poor iron removal, non-optimal pH, and inefficient construction. Figure 5 shows the percent As and Fe removals as a function of the initial Fe/As ratio, and Figure 6a indicates the As removed as a function of the initial total As concentration in the IRU. Figure 6b shows the percent As removed as a function of the ratio As(III)/Total-As in the IRU. The obtained results are in accordance with the literature that the %As removal increases with an increase in %Fe removal and the Fe/As ratio, and decreases with an increase in Total As and in As(III)/Total-As ratio. Figure 6a also shows a good correlation between percent As removed compared to initial arsenic concentration with an r2 value of 0.83. When calculating this r2 value the points encircled are shown in the figure, but not taken into consideration since the points show that

100 90 80

% Removal

70 60 50 40 30

Fe

20

As

10 0 0

Figure 5.

20

60 80 Fe/As ratio

40

100

120

140

Removal of Fe and As as a function of the initial Fe/As ratio (M/M) in IRU.

b

0.60

0.60

0.50

0.50

As(III)/Total Arsenic

Initial Arsenic concentration mg/L

a

0.40 0.30 y = -0.0054x + 0.68 R2 = 0.83

0.20 0.10

0.40 0.30 0.20

y = -0.0095x + 0.98 R2 = 0.41

0.10 0.00

0.00 0

40 80 20 60 % Arsenic removed

0

100

20 40 60 80 % Arsenic removed

100

Figure 6. Percent arsenic removed with respect to (a) initial As concentration, and (b) As(III)/Total-As in the IRU. * denotes the omitted results for calculation of regression coefficient.

268

even at low Fe/As ratios and high initial As concentration higher As removals were obtained. At one of those IRU already after aeration 34% As was removed which was not the case at the other IRU. One of the reasons behind this could be that the groundwater contain other ions which may promote the oxidation of As(III), but since other ions are not measured it is difficult to comment on this finding. 3.3

Other factors affecting the removal

The sedimentation experiments conducted in West Bengal showed that As removal increased with time (Fig. 7) and that at high initial As concentrations the effect of sedimentation time was high, while at low initial As concentrations the difference in As removal after 8 hours was very low compared to 24 hours removal. The results also showed that there was no additional As removal after 24 hours of sedimentation compared to 1 month sedimentation (results are not shown in the figure). The Fe/As ratio was in the range of 7–17 (M/M), whereas the respective phosphate and silicate concentrations were in the range of 0.6–0.9 mg/L and 17–18 mg/L. The pH was in the range of 6.99–7.13. The sedimentation time allowed in case of Bangladesh in the plain sedimentation experiments was 6–8 days. Since a high Fe/As ratio was present it can be assumed based on the results from West Bengal that sufficient time is allowed to reach the equilibrium concentration. The time of sedimentation allowed in the IRU depends on the amount of water withdrawal and is generally less than 24 hours. Similar residual As concentrations in treated water with IRU compared to treated water with plain sedimentation may indicate that sufficient time is allowed for treatment in the IRU. Wilkie & Hering (1996) observed negligible pH effects on As(V) adsorption on hydrous ferric oxide in the pH range 6–7, while a decrease in As(V) adsorption on hydrous ferric oxides was observed with an increase in pH above 7, and an increase in pH from 7 to 8 resulted in a decrease in adsorption by 20% (Pierce & Moore 1982, Wilkie & Hering 1996). The pH in the raw water in the IRU ranged from 6.5–7 and an increase in pH was observed after treatment with resulting pH values in the range of 6.88–7.84. One of the ways of comparing the present results with Wilkie & Hering (1996) is to compare the ratio of amount of As removed with the amount of Fe removed as a function of pH (Fig. 8). The results show that higher As/Fe removal ratios were obtained at lower pH except for two IRU. At these 2 IRU very high As/Fe removal

Arsenic concentration (mg/L)

0.80 0.70 0.60

7

0.50

17 11

0.40

10

0.30

8

0.20 0.10 0.00 0

5

10

15 Time (hours)

20

25

30

Figure 7. As removal in buckets with time in water from West Bengal. The numbers 7, 17, 11, 10 and 8 in the legend indicates the Fe/As ratio (M/M) present in the raw water.

269

As removed/Fe removed (mg/mg)

0.08 0.07 0.06 0.05 0.04 0.03 0.02 0.01 0 6.8

7

7.2

7.4 pH

7.6

7.8

8

Figure 8. Effect of equilibrium pH on the As removed vs Fe removed. * denotes the omitted results for calculation of regression coefficient in Figure 6.

ratios were observed. These are the same IRU as shown in Figure 6a where higher As removals were observed even at lower Fe concentrations.

4

CONCLUSIONS

The present investigations showed that IRU are also capable of removing As. The results showed however that the present IRU are not working properly and hence the resulting Fe concentration in treated water is above 1 mg/L. The major reason is shortcutting the aerated water directly to the treated water tank without any sedimentation or filtration due to improper construction. By improving the Fe removal capacity of the IRU the As removal capacity is also increased. The results indicate that apparently there is no significant difference between removal of As using IRU, or with simple plain sedimentation in buckets, but there is a possibility of better performance if the IRU are improved. In case of plain sedimentation care has to be taken not to disturb the water during collecting the clean top layer, free from both Fe and As. Utilising plain sedimentation requires storage space for at least one day’s consumption. The IRU are expensive compared to plain sedimentation, but cheap compared to other treatment options available at present. They can be constructed with local material and maintained with local skills. The As removal capacity of the IRU can be increased if pre-oxidation of As(III) to As(V) can be achieved with addition of oxidation agents like KMnO4. The main drawback of using IRU for As removals is sludge production. Based on the present findings of a necessary Fe/As ratio of 80 (M/M) for removal of As below 50 ␮g/L approximately a quarter of the tube wells exceeding As concentration of 50 ␮g/L can be treated to meet the Bangladesh standards by using naturally occurring Fe for the removal of As.

ACKNOWLEDGEMENTS The authors are grateful to Dr. Jens Thøgersen and to the staff of DPHE-DANIDA, as well as UWASP (Noakhali), Bangladesh for the assistance received. Furthermore, the authors express their sincere thanks to the help of members of BSSKS/RCS, West Bengal, India. The authors 270

would like to thank Gunnar Jacks and Prosun Bhattacharya for their valuable comments on an earlier draft of the manuscript. REFERENCES Ahmed K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A.H., Imam, M.B., Khan, A.A. & Sracek, O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh: an overview. Applied Geochemistry 19: 181–200. BGS 1998 DPHE/BGS National Hydrochemical Survey. http://www.bgs.ac.uk/arsenic/bangladesh/datadownload.htm (Accessed on September 21, 2001). Borho, M. & Widerer, P. 1996. Optimized Removal of Arsenate(III) by Adaptation of Oxidation and Precipitation Processes to the Filtration Step Water, Science and Technology 34(9): 25–31. Gao, Y. & Mucci, A. 2001. Acid base reactions, phosphate and arsenate complexation, and their competitive adsorption at the surface of goethite in 0.7 M NaCl solution. Geochimica et cosmochimica Acta 65(14): 2361–2378. Hering, J. G., Chen, P.-Y., Wilkie, J. A., Elimelech, M. & Liang, S. 1996. Arsenic Removal by Ferric Chloride, Journal AWWA: 155–167. HMSO 1972. Analysis of Raw Water, Portable and Waste Waters. London, HMSO. Hossain, D. Md. & Ali, A. M. 1997. Arsenic Removal from Groundwater by Co-precipitation, Paper No. 19, Training Course on Arsenic Problems, & Dearsination of Drinking water for use in Bangladesh, International Training Network-Bangladesh. Hug, S.J., Cononica, L., Wegelin, M., Gechter, D. & Gunten, U.V. 2001. Solar oxidation and removal of arsenic at circumneutral pH in iron containing waters. Environmental Science and Technology 35: 2114–2121. Jain, A. & Loeppert, H. 2000. Effect of Competing Anions on the Adsorption of Arsenate and Arsenite by Ferrihydrite. J. Environ. Qual. 29: 1422–1430. Mamtaz, R. & Bache, D.H. 2000. Low-Cost seperation of Arsenic from water: with special reference to Bangladesh. J. CIWEM 14: 260–269. Manning, B.A. & Goldberg, S. 1996. Modeling Competetive Adsorption of Arsenate with Phosphate and Molybdate on Oxide Minerals. Soil Sci. Soc. Am. J. 60: 121–131. Meng, X. & Korfiatis, G.P. 2000. Effects of silicate, sulfate and carbonate on arsenic removal by ferric chloride. Wat. Res. 34: 1255–1261. Meng, X., Korfiatis, G.P., Christodoulatos, C. & Bang, S. 2001. Treatment of Arsenic in Bangladesh well water using a household co-precipitation and filtration system. Wat. Res. 35: 2805–2810. Nickson, R.T., McARthur, J.M., Ravenscroft, P., Burgess, W.G. & Ahmed, K.M. 1998. Mechanism of arsenic release to groundwater, Bangladesh and West Bengal. Applied Geochemistry 15(4): 403–413. Pierce, M.L. & Moore, C.B. 1982. Adsorption of Arsenite and Arsenate on Amorphous Iron Hydroxide, Water Res. 16: 1247–1253. Raven, K.P., Jain, A. & Loeppert, R.H. 1998. Arsenite and arsenate adsorption of Ferrihydrite: Kinetics, Equilibrium, and adsorption envelopes. Environmental Science and Technology 32: 344–349. Smedley, P.L. & Kinniburgh, D.G. 2002. A review of the source behaviour and distribution of arsenic in natural waters. Applied Geochemistry 17: 517–568 American Public Health Association (APHA) 1995a. Standard Methods for the Examination of Water and Wastewater, American Public Health Association, 1015 Fifteenth Street, NW Washington, DC: 3.49–3.51. American Public Health Association (APHA) 1995b. Standard Methods for the Examination of Water and Wastewater, American Public Health Association, 1015 Fifteenth Street, NW Washington, DC: 3.32–3.33. American Public Health Association (APHA) 1995c. Standard Methods for the Examination of Water and Wastewater, American Public Health Association, 1015 Fifteenth Street, NW Washington, DC: 4.114–4.115. American Public Health Association (APHA) 1995d. Standard Methods for the Examination of Water and Wastewater 1995d. American Public Health Association, 1015 Fifteenth Street, NW Washington, DC: 3.13–3.18. Su, C. & Puls, R.W. 2001. Arsenate and arsenite removal by zerovalent iron: Effects of phosphate, silicate, carbonate, borate, sulfate, chromate, molybdate, and nitrate relative to chloride. Environmental Science and Technology 35: 4562–4568. WHO 1996. Guidelines For Drinking Water Quality, Recommendations, 2. Edition, Vol. 2, World Health Organisation. Wilkie, J.A. & Hering, J.G. 1996. Adsorption of arsenic onto hydrous ferric oxide: effects of adsorbate/adsorbent ratios and co-occurring solutes. Colloids and surfaces A 107: 97–110.

271

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

A simple and environmentally safe process for arsenic remediation – laboratory and field evaluation Kshipra Misra, M.T. Companywala, Sanskriti Sharma, Alips Srivastava & P.C. Deb Naval Materials Research Laboratory (NMRL), DRDO, Ministry of Defence, Addl. Ambernath, India

ABSTRACT: This paper reports the results of laboratory and field evaluation of a simple and environment-friendly arsenic removal filter. The filter works on the simple principle of co-precipitation and adsorption followed by filtration through treated sand. An easily available processed waste of steel industry is used as a reactive medium in the filter. Laboratory trials of the filter have successfully been completed. Prototype filters are installed for field trials in the arsenic-affected villages of West Bengal and have been reported to be successfully operating. Salient features of the filter include its cost-effectiveness, and easy operation and maintenance, involving only normal washing and replacement of the media. It is suitable for household use and requires no energy source. The waste generated can be converted into non-leachable cement matrix of M-25 standard grade impermeable concrete blocks used in construction industry. This makes the system ecofriendly. The unit flow rate of capacity 15 L/h and 30 L/h filtered water and quality conform to the drinking water standards for arsenic and iron.

1

INTRODUCTION

An alarmingly large population of India and Bangladesh, 66 million in the Gangetic belt of India and 79.9 million in Bangladesh (Bose & Sharma 2002, Ahmed et al. 2004) is exposed to arsenic poisoning due to continuous usage of arsenic-contaminated ground water. Arsenic concentration in the water of these regions above the permissible limit (Chakraborti et al. 2003, USEPA 2001, 2004). Arsenic contamination of groundwater in these areas has mainly occurred due to natural reasons (Bose & Sharma 2002, Ahmed et al. 2004). According to the most accepted and most plausible theory, in the Late Pleistocene/ Holocene period, iron and arsenic-bearing minerals in upstream of the Ganges river belt may have undergone oxidation due to exposure to atmosphere during erosion, resulting in subsequent mobilization of arsenic and iron downstream. The mobilized iron got precipitated as iron oxy-hydroxide and arsenic got either adsorbed onto or co-precipitated with iron oxy-hydroxide. These arsenic containing precipitates then got deposited in the Gangetic delta region in the form of iron oxy-hydroxide coating on aquifer sediments. In the present day situation, reducing conditions prevailing in the sub-surface environment is causing dissolution of this coating and mobilization of adsorbed/co-precipitated arsenic (Bhattacharya et al. 1997, Nickson et al. 1998, 2000, McArthur et al. 2001. Most of the affected people in the subcontinent are poor villagers and so the commonly available expensive technologies become economically non-viable. More so, the delicacy of these technologies and the subsequent operation and maintenance (Zaw et al. 2002, Katsoyiannis et al. 2002) add to their expenses, apart from being inconvenient to be used by villagers. Considering that a large population at risk, there is an urgent need to develop ways to mitigate this problem by reducing the level of arsenic in drinking water to tolerable limits through easy and inexpensive means. The arsenic removal filter reported here is designed to provide a low cost, easily available, eco-friendly arsenic removal system for the rural people. 273

2

MATERIALS AND METHODS

The reactant material (media), a processed waste from Steel industry, has been obtained from M/s Tata Wires Ltd., Mumbai. Sand used has been obtained from the riverbank of River Yamuna in Delhi, India, and from the riverbank of River Ganga in Kolkata, India. Fine cloth filter has been procured from a local cloth merchant in Mumbai. AR quality reagents and Milli-Q grade water have been used for solution preparation. Solutions of As⫹3 and As⫹5 have been prepared using corresponding salts, NaAsO2 and Na2HAsO4 ⭈ 7H2O, respectively. Mixture of As⫹3 and As⫹5 (in the ratio of 1:1) has been prepared by dissolving equimolar amount of corresponding salt in Milli-Q grade water. The reactant material is soaked overnight in water before using in the filter. The sand used is subjected to physical treatment (washing and heat treatment) prior to using it in the arsenic removal filter. The reactant material and sand have been characterized for their surface area and composition using Micromeritics ASAP 2010 Surface Area Analyzer at Centre for Fire, Environment and Explosives Safety (CFEES), Delhi, and by Phillips X-ray fluorescence (XRF) at Durgapur Steel Plant, Durgapur, respectively. The surface texture of the reactant material was carried out on Scanning Electron Microscope (Model: Leo 1455). 2.1 Metal analysis The variation in the pH of pure water and of arsenic solution when allowed to percolate down through the reactant material and through sand has been determined using a pH meter (Model: Elico LI-120). Arsenic concentration in water, prior to and after treatment, has been measured as per ASTM method (ASTM D 2972-88) using Hydride Generator (Model: HG-3000) attached to AAS (Model: GBC 904AA) at Centre for Fire, Environment and Explosives Safety (CFEES), Delhi, India. Iron concentration was also determined using same AAS. 2.2 Design of arsenic removal filter Arsenic removal filter (Fig. 1) has been designed and fabricated both in plastic and in stainless steel. Two filter systems have been designed and evaluated, one operating at a flow rate of 15 L/h and the other operating at a flow rate of 30 L/h. 2.3 Waste disposal Although the waste generated during arsenic removal process is not environmentally harmful as such, as reported by earlier workers, yet disposal of arsenic-laden waste is an important aspect under growing environmental regulations. Therefore, precipitate formed during reaction and the used sand is being disposed off in the form of impermeable concrete blocks of M-25 (Singh, 1982) standard grade used in construction industry resulting in no waste generation in the process and making the technology environment-friendly and green.

3

RESULTS AND DISCUSSION

3.1 Characteristics of the reactant material Scanning electron micrograph (Fig. 2) of the material at 500 times magnification shows the fibrous elongated morphology. Characteristics of the reactant material and sand (Table 2) clearly indicate that the reactant material is nothing but 99% iron and acts as zero-valent iron. Surface area value of sand indicates that its adsorption capacity is low and is basically functioning as a fine filter in this process. The arsenic removal filter, therefore, works on the simple principle of co-precipitation of arsenic with iron and adsorption of this precipitate on iron oxyhydroxides (Su et al. 2001, Manning et al. 274

Figure 1. Schematic diagram of arsenic removal filter. Explanations: 1: Inlet for arsenic contaminated water; 2: Reactant material; 3: Fine cloth filter; 4: Treated sand; 5: Fine cloth filter; 6: Arsenic-free water; 7: Outlet for arsenic-free water; 8: Container for arsenic-free water; 9: Container for treated sand; and 10: Container for reactant material. Table 1.

A comparison of the arsenic removal filter systems.

Specification

System I

System II

Flow rate Amount of treated sand Reactant (Steel plant waste) Initial as concentration Final as concentration Volume of treated water Quality of water for drinking Leaching of other metals

15 L/h 1500 g 500 g 1 mg/L ⬍3 ␮g/L 750 L Suitable None

30 L/h 3000 g 1000 g 1 mg/L ⬍3 ␮g/L 1750 L Suitable None

2002, Melitas et al. 2002, Nikolaidis et al. 1998), followed by filtration through treated sand. Probable reactions involved in the process are as given below: (1)

(2)

Sodium salts of arsenite and arsenate get ionized in aqueous solution. Both arsenite and arsenate oxyanions are removed further by co-precipitation (as FeAsO4 and FeAsO3) and by adsorption onto ferric oxyhydroxide solids. The same has been reported by a number of workers earlier also (Su et al. 2001). 275

Figure 2. Table 2.

SEM micrograph of reactant material (⫻500). Characteristics of sand and reactant material. Surface area (BET) (m2/g)

Adsorbent

pH (Water)

pH (As solution)

Fe (%)

Al (%)

Mn (%)

Sand (Yamuna) Sand (Ganga) Reactant material

10.2–10.5

10.2–10.5

8.9–10.5

10.5–11.0

Not detected 79.2–80.0

1

8.3–8.5

7.5–8.0

4.7–5.0

11.5–12.0

Not detected 79.8–80.0

4

8.5–9.0

8.8–9.0

99.2–99.5

Not detected

0.42–0.45

0.5

3.2

Si (%)

Trace amount

Laboratory evaluation

3.2.1 Optimization of flow rate Keeping the amount of reactant material and treated sand constant, 500 g and 1500 g respectively, experiments have been carried out to study the effect of flow rate of arsenic contaminated water (As⫹3 or As⫹5 or 1:1 mixture of As⫹3 and As⫹5) through the filter on the removal efficiency of the filter. The results of these experiments show that irrespective of the arsenic species present in water, effective removal of arsenic can be achieved up to a maximum flow rate of 15 L/h in first system. Arsenic concentration in filtered water increases above prescribed limits as the flow rate exceeds this value. However, it has also been established that if the amount of reactant material and treated sand was raised to 1000 g and 3000 g, respectively, maximum allowable flow rate that could be achieved is 30 L/h, by changing the dimensions of the filter accordingly as already explained in the experimental section. The results are depicted in Figure 3. 3.2.2 Effect of initial arsenic concentration The effect of initial arsenic (1:1 mixture of As⫹3 and As⫹5) concentration (varying from 1–4 mg/L) on the arsenic removal efficiency of the filter, in terms of total volume of water filtered (final 276

Final As conc. (µg/L)

Initial As conc. ~ 1 mg/L Amount of Adsorbent = 500 g Amount of Treated Sand = 1500 g

28

Initial As conc. ~ 1 mg/L Amount of Adsorbent = 1000 g Amount of Treated Sand = 3000 g

24 20 16 WHO / EPA Drinking Water Limit

12 8 4 0 0

5

10

15

20

25

30

35

40

45

50

55

Flow Rate (Lph)

Figure 3.

Optimization of flow rate. 1750

Volume of Treated Water (L)

1800 1600

15 Lph 30 Lph

1455

1400 1165

1200 1000 800

875 750 625 500

600

375

400 200 0 1

2

3

4

Initial As conc. (mg/L)

Figure 4.

Effect of initial arsenic concentration on treated water volume at the two different flow rates.

arsenic concentration in filtered water ⬍10 (g/L), using optimized amounts of reactant material and treated sand for the two flow rate systems has been studied and is shown in Fig. 4. As expected, an increase in the arsenic concentration in water leads to a decrease in the total volume of water that can be treated using this filter. 3.2.3 Water quality The filtered water collected in the third chamber has been analyzed for its arsenic concentration, iron (that may leach out from the reactant material during the process) and microbes. The results as enlisted in Table 3 clearly indicate that the quality of the filtered water conforms to the internationally (WHO and US EPA) set drinking water standards (USEPA, 2004). 3.3

Field evaluation

After successful laboratory evaluation, seven filters of 15 L/h flow rate were installed in four villages, namely, Kamdevkati, Raghavpur, Simulpur and Chatra villages of 24 Paraganas (N) district of West Bengal, India, to test the viability of this technology in field conditions (Table 4). 277

Table 3.

Results of water analyses.

Type of As-species As(III) As(V) Mixture of As (III) and As (V) in the ratio of 1:1 Table 4.

As conc. (␮g/L)

Fe conc. (mg/L)

E.coli (count/100 mL) after 48 hrs.

Initial

After treatment

Initial

After treatment

Initial

After treatment

1000 1050

⬍.3 ⬍.3

Not detected Not detected

⬍0.3 ⬍0.3

8 8

0 0

1025

⬍.3

Not detected

⬍0.3

8

0

Field evaluation data.

Date of installation Site of installation 23/09/03 07/10/03 20/11/03 20/11/03 20/11/03 21/11/03

Capacity (mg/g)

21/11/03

Kamdevkati Village (Stainless Steel Kit) Chatra Village (Plastic Kit)2 Kamdevkati Village (Stainless Steel Kit) Kamdevkati Village (Plastic Kit) Simulpur Village Raghavpur Village (Stainless Steel Kit) Chatra Village (Plastic Kit)3

Total volume Iron concentration (mg/L) of water filtered1 Initial Final

Initial

Final

10,050 L 0.040 ⫾ 0.001 0.042 ⫾ 0.001 0.068 ⫾ 0.01 1400 L 0.168 ⫾ 0.02 0.063 ⫾ 0.001 0.271 ⫾ 0.02

0.004 ⫾ 0.001 0.003 ⫾ 0.001

7260 L

0.105 ⫾ 0.02 0.210 ⫾ 0.02

0.049 ⫾ 0.001

6860 L 6960 L

0.084 ⫾ 0.001 0.126 ⫾ 0.02 0.462 ⫾ 0.02 0.168 ⫾ 0.02

0.135 ⫾ 0.02 ⬍0.003 ⫾ 0.001 0.374 ⫾ 0.02 0.005 ⫾ 0.001

7110 L

0.168 ⫾ 0.02 0.189 ⫾ 0.02

0.180 ⫾ 0.02 ⬍0.003 ⫾ 0.001

2060 L

0.168 ⫾ 0.02 0.105 ⫾ 0.02

0.271 ⫾ 0.02

a

4

Arsenic concentration (mg/L)

0.004 ⫾ 0.001

b

6

3

0.003 ⫾ 0.001

4

2

2

1

0

0 0

2 4 As conc. (mg/L)

6

0

0,2 As conc. (mg/L)

0,4

Figure 5. Comparison between the capacity of the reactant material used in the filter for the removal of arsenic based on the results obtained from the tests in the laboratory (a) and field evaluation (b).

3.4

Comparison of laboratory and field data

A very good concordance was observed between laboratory and field results (Fig. 5) in terms of the capacity of the reactant material used in the filter for the removal of arsenic from the water. The capacity is calculated on the basis of initial arsenic concentration in influent water with respect to the total quantity of water filtered by filter in the laboratory and in field so far. The results as discussed above clearly indicate that the quantity of reactant material used in the filter can be easily and more efficiently used for initial higher concentration of arsenic in water, i.e., up to 4 mg/L and thereafter it remains constant. Therefore, the filter can be successfully used 278

Table 5. Results of leaching tests on the wastes generated during arsenic removal processes. Arsenic concentration in filtered water (␮g/L) Type of waste

Laboratory samples

Field samples

Waste Concrete blocks

BDL* BDL

BDL BDL

* BDL: Below Detection Limit i.e. ⬍3 ␮g/L.

up to a maximum initial concentration of 4 mg/L keeping all the experimental conditions same as discussed in experimental section. 3.5

Leaching tests for waste and concrete blocks

Leaching tests carried out for the waste generated during the process as well as for concrete blocks as per the standard Toxicity Characteristic Leaching Procedure (TCLP) for solid wastes (EPA protocol SW-846-1311) (www.iwrc.org), gave results that are tabulated in Table 5.

4

CONCLUSIONS

The reports collected from the field trials indicate commendable performance of the filters. However, it has been observed that stainless steel filters are more durable than plastic filters for long-term usage and are recommended for further use. The water filter for arsenic removal as discussed above can provide a reliable solution to the basic problem of arsenic contamination in ground water because of its following features:

• • • • • •

Requires no energy sources Easy maintenance Cost-effective Environment-friendly User Friendly Easy waste disposal

ACKNOWLEDGEMENTS Authors wish to express their sincere gratitude to Dr. J.N. Das, Director, NMRL, Ambernath, for granting permission to publish this work. Authors also wish to acknowledge the help provided by Mr. P.K. Singh, NMRL, for carrying out SEM analysis of the samples and Mr. Rajeev Goel, CFEES, Delhi for carrying out arsenic analysis of water samples. Authors would also like to thank Prosun Bhattacharya at the Royal Institute of Technology, Stockholm, Sweden and K:M: Ahmed from the University of Dhaka, Bangladesh for their constructive suggestions on an earlier draft of the manuscript.

REFERENCES Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A.H., Imam, M.B., Khan, A.A. & Sracek, O. 2004. Arsenic contamination in groundwater of alluvial aquifers in Bangladesh: An overview. Applied Geochemistry 19(2): 181–200.

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Berg, M., Tran, H.C., Nguyen, T.C., Pham, M.V., Schertenleib, R. & Giger, W. 2001. Arsenic contamination of groundwater and drinking water in Vietnam: A human health threat. Environ. Sci. Technol. 35 (13): 2621–2626. Bhattacharya, P., Chatterjee, D. & Jacks, G. 1997. Occurrence of Arsenic-contaminated Groundwater in Alluvial Aquifers from Delta Plains, Eastern India: Options for Safe Drinking Water Supply. Water Resources Development 13(1): 79–92. Bose, P. & Sharma, A. 2002. Role of iron in controlling speciation and mobilization of arsenic in subsurface environment. Water Research 36: 4916–4926. Chakraborti, D., Mukherjee, S.C., Pati, S., Sengupta, M.K., Rahman, M.M., Chowdhury, U.K., Lodh, D., Chanda, C.R., Chakraborti, A.K. & Basu, G.K. 2003. Arsenic groundwater contamination in Middle Ganga Plain, Bihar, India: A Future Danger? Environmental Health Perspectives 111(9): 1194–1200. IWRC 2004. Toxicity characteristic leaching procedure testing parameters. Iowa Waste Reduction Center/ University of Northern Iowa URL: http://www.iwrc.org/summaries/TCLP.cfm (Accessed on September 6, 2004. Katsoyiannis, I.A. & Zouboulis, A.I. 2002. Removal of arsenic from contaminated water sources by sorption onto iron-oxide-coated polymeric materials. Water Research 36: 5141–5155. Manning, B.A., Hunt, M.L., Amrhein, C. & Yarmoff, J.A. 2002. Arsenic(III) and Arsenic(V) Reactions with Zerovalent Iron Corrosion Products, Environ. Sci. Technol. 36, 5455–5461. McArthur, J.M., Ravenscroft, P., Safiullah, S. & Thirlwall, M.F. 2001. Arsenic in groundwater: Testing pollution mechanism for sedimentary aquifers in Bangladesh. Water Resour. Res. 37 (1) 109–118. Melitas, N., Wang, J., Conklin, M., O’Day, P. & Farrell, J. 2002. Understanding Soluble Arsenate Removal Kinetics by Zerovalent Iron Media, Environ. Sci. Technol. 36: 2074–2081. Nickson, R.T., McArthur, J.M., Burgess, W.G., Ahmed, K.M., Ravenscroft, P. & Rahman, M. 1998. Arsenic poisoning of Bangladesh groundwater. Nature 395: 338. Nickson, R.T., McArthur, J.M., Ravenscroft, P., Burgess, W.G. & Ahmed, K.M. 2000. Mechanism of arsenic release to groundwater, Bangladesh and West Bengal, Appl. Geochem. 15(4): 403–413. Singh G. 1982. Hand Book of Civil Engineering, Part II, pp. 29. Su, C. & Puls, R.W. 2001. Arsenate and Arsenite removal by zerovalent iron: Kinetics, redox transformation, and implications for in-situ groundwater remediation, Environ. Sci. Technol. 35: 1487–1492. Nikolaidis, N.P., Lackovic, J. & Dobbs, G. 1998. Arsenic Remediation Technology-AsRT. U.S. Application #60/050,250 (Patent Pending). Environmental Research Institute, University of Connecticut. URL: http://www.eng2.uconn.edu/⬃nikos/asrt-brochure.html. USEPA 2001. National Primary Drinking Water Regulations; Arsenic and Clarifications to Compliance and New Source Contaminants Monitoring; Final Rule, Federal Register 66 (14), 6975. USEPA 2004. List of Drinking Water Contaminants and MCLs. United States Environmental Protection Agency. URL: http://www.epa.gov/safewater/mcl.html (Accessed on September 6, 2004). Zaw, M. & Emett, M.T. 2002. Arsenic removal from water using advanced oxidation process. Toxicology Letters 133(1): 113–118.

280

Section 5: Management of arsenic-rich groundwaters

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Management of the groundwater arsenic disaster in Bangladesh K.M. Ahmed Department of Geology, University of Dhaka, Dhaka, Bangladesh

ABSTRACT: Arsenic contamination of groundwater in Bangladesh has emerged as the largest water pollution event in the world. In Bangladesh it had been thought that more than 97% of the population had access to safe drinking water until recently and because of detection of arsenic in groundwater, the main source of drinking water in rural and urban areas, the access has come down to 80%. The recent increase in rice production is also attributed to groundwater irrigation and about 70% of irrigation water is abstracted from the aquifers, which account for 85% of the total abstracted groundwater. The presence of arsenic in groundwater has become a major issue of management of both drinking and irrigation water in the country. Analysis of about 4 million wells by field kit show that about one third of the wells exceed the BDWS of 50 ␮g/L. If the WHO provisional guideline value, 10 ␮g/L, is considered two thirds of the wells become unsafe. About 30 to 70 million people are exposed to unsafe level of arsenic in their drinking water. Yet another potential intake source remains unknown i.e. the entries through the food chain due to irrigation with arsenic contaminated groundwater. Scientific investigations have demonstrated that arsenic contamination does not occur randomly; rather geology and hydrogeology control it. Groundwater in Bangladesh has never been considered as a precious resource; rather it has been used indiscriminately. For management of arsenic in groundwater, a pragmatic mitigation policy is needed with a holistic approach. Legal and institutional reforms are necessary to address the issue. Concepts of safe drinking water and integrated water resources management should be adopted in order to manage the problem scientifically. Proper assessment has to be performed to avoid substitution of new risks while implementing new sources. Community participation has to be ensured in the management by raising the level of awareness about the quality and quantity of the vital resource.

1

INTRODUCTION

Bangladesh relies intensely on groundwater for rural and urban water supply and achieved remarkable success in providing access to safe water to 97% of the population. Increased irrigation coverage with groundwater contributing for more than 70% made the country safe reliant in rice production, the staple food the 130 million people. The exponential increase in groundwater exploitation has been prompted by easy availability and low cost technologies. The detection of arsenic above admissible limits in shallow groundwater of Bangladesh has emerged as a severe environmental hazard and the use of groundwater both for drinking and irrigation purposes being questioned. Groundwater management is in poor state despite large dependence on groundwater. The theme of this paper is management of groundwater for safe and sustainable use, mainly for drinking purposes as it warrants the top priority among all uses. 1.1

Arsenic contamination situation

Since the first detection of arsenic in 1993, a number of studies have been carried out to determine the extent of the problem. Until recently a large proportion of the country’s 8 to 10 million water supply wells have been tested using field kits. At the same time, water samples from a relatively 283

small number of wells have also been analyzed using laboratory techniques. Field kit analysis of about 4 million wells from 402 Upazilas (sub-district) out of about 500 reveal that about 30% of the wells have arsenic above the Bangladesh drinking water standard (BDWS) of 0.05 mg/L (Fig. 1a). On the other hand, laboratory analyses of 44000 wells by various agencies show that 34% wells yield water with more than 0.05 mg/L As (Fig. 1b). Among the tested Upazilas there are extreme variations in the extent of occurrence of wells with As above the BDWS – in some Upazilas almost all tested wells exceed the limit, in some 0.05 mg/L

% Tested Wells

120 100 80 60 40 20 0 BAMWSP UNICEF & Full 210 DPHE 192 Total 402 156 Others 54 Screened Upazila

% of Tested Wells

50 40 30 20 10 0 < 0.01 0.01-0.05 > 0.05 Arsenic Concentrations (mg/L) 0

20

% of wells 40 60

% of wells

=150 to < 250 >=250 to =500 to =750 to =1000 >0.05 mg/L

Figure 1. (a) Percentage of wells exceeding BDWS under various field kit surveys (summary of about 4 million tests); (b) Percentage of wells under different concentrations ranges (summary of 44,000 laboratory analyses); (c) BAMWSP data from 66 upazila show that a larger proportion of wells from depth ⬎500 feet exceed the BDWS (Fig. 1c).

284

other Upazilas none exceed the limit. However, there are distinct spatial patterns in occurrence as shown in Figure 2. It is evident from the figure that significantly large proportions of wells located in the southern sub-districts exceed the BDWS compared to the wells located in the north. It has been reported by number of studies that spatial distribution of As occurrences is controlled by surface geology, i.e. most wells located in the areas occupied by the fine and young floodplains and deltaic sediments exceed BDWS more frequently, wells installed in Holocene coarse fan deposits rarely exceed the BDWS and wells developed in Pleistocene and older never exceed the BDWS (BGS & DPHE 2001, Ahmed et al. 2004). Also there is depth control in the occurrence of high As in groundwater. The peak concentration occurs at 20–40 m, whereas aquifers above and below have lower concentrations. BGS & DPHE (2001) reports that aquifers deeper than 150 m have consistently low arsenic. Other studies report occurrence of low As water at shallower (van Geen et al. 2004) and deeper depths (JICA 2003). However, BAMWSP data from 66 upazila show that a larger proportion of wells from depth ⬎500 feet exceed the BDWS (Fig. 1c). These findings are contradictory to earlier studies and needs further investigation. It has been established from different studies 88°

89°

90°

91°

92°

26°

22°

22°

23°

23°

24°

24°

25°

25°

26°

Rivers % of wells exceding 0.05 mg/L 0-1 1.1 - 20 20.1 - 40 40.1 - 60 60.1 - 80 80.1 - 100 No data

21°

21°

Bay of Bengal

80

88°

89°

0

80 kilometers

90°

91°

92°

Figure 2. Distribution of wells exceeding the BDWS for arsenic in different Upazilas (data from BAMWSP, DPHE/UNICEF).

285

40.08 2003

25.04 74.96 38.5

2002

2001

24.67 75.33 37.66 23.68 76.32 35.57

2000

24.90 75.10

Groundwater (%)

Figure 3.

Surface Water (%)

Total Area (Million Hectres)

Area covered by surface and groundwater irrigation (data from BADC irrigation census).

that there is no specific depth for the occurrence of As safe water, rather it is controlled by the subsurface geology and hydrologic conditions (Ahmed 2003a). 1.2

Share of groundwater in water supply and irrigation

Groundwater has been the main element for two recent achievements of Bangladesh in the field of access to safe water and food security. Due to extensive use of groundwater, facilitated by easy availability of prolific aquifers, low-tech installation procedure and affordable cost, 97% of the total population came under the safe water supply. The number of domestic water supply wells increased many folds over last 3 decades and 90% of these are privately owned (van Geen et al. 2002). Groundwater is also the main sources of municipal water supplies in urban areas including the capital city Dhaka. In fact Dhaka is one of the mega cities of the world, which rely almost entirely on groundwater for water supply (Ahmed et al. 1999, Morris et al. 2003). The second success that the country achieved recently is attaining self-sufficiency in rice production. This has been made possible by the role of groundwater-based minor irrigation systems, which now accounts for more than 75% of the total coverage. Though irrigation started in Bangladesh by using surface water, the source has shifted from surface to groundwater as shown in Figure 3. Both the successes have been posed with the recent threat due to As occurrences in groundwater. As nearly 30% of the wells exceed the BDWS, the population previously considered to have access to safe water is now known to be exposed to high As in their drinking water. Also the issue of arsenic transfer through food chain due to irrigation with high As water is becoming a matter of concern as many studies have reported arsenic buildup in soil and crops (Huq & Naidu 2003, Farid et al. 2003). 1.3

Arsenic management issue

Occurrences of As above permissible limit have exposed millions of people to mass poisoning (Smith et al. 2000). It is considered that the groundwater As catastrophe has emerged due to poor or no management of drinking water sources in Bangladesh (Ahmed & Ravenscroft 2000). For the last 5–6 years, there have been efforts to mitigate the problem but the pace of mitigation activities does not match the extent and severity of the problem (Chakraborty et al. 2002). Various options have been suggested and tested at various locations (van Geen et al. 2002, van Geen et al. 2003). There are suggestions for a variety of mitigation options and strategies (WHO 2000, Hoque et al. 2000, GOB 2002, Yu et al. 2003, Alaerts & Khouri 2004). At the same time there are uncertainties regarding various options which need further investigations before being applied (Burgess et al. 2002a, b, Cuthbert et al. 2002, Caldwell et al. 2003). Also the role of irrigation in the As mobilization process is considered important (Harvey et al. 2002) and the food chain issues are becoming 286

more and more important (Huq & Naidu 2003). Therefore, one needs to consider both the water supply and irrigation issues together while planning for As management in Bangladesh. The current paper attempts to provide a preliminary risk assessment of different groundwater-based As-safe sources based on limited data. At the same time a broader framework for groundwater management in the country will be provided under the existing guidelines and policies. Also research needs and local capacity building issues will be highlighted for sustainable management of groundwater. 1.4

Hydrogeological setting of Bangladesh

Both the spatial and depth distribution of As occurrences in Bangladesh groundwater is controlled by geological and hydrogeological factors. It is therefore important to provide a broad overview of regional hydrogeology and groundwater occurrences. The country can be broadly divided into six major hydrogeological units (Ahmed 2003b). However, if minor details are considered the number of units increases up to 40 (UNDP 1982, MPO 1985). Figure 4 shows the major hydrogeological zones of the country which are: Zone I – Holocene Piedmont Plains, Zone II – Holocene Deltaicand Flood-plains, Zone III – Pleistocene Terraces, Zone IV – Holocene Depressions, Zone V – Tertiary Hills, and Zone VI – Holocene Coastal Plains. Aquifer conditions and quality of groundwater vary significantly from unit to unit. Table 1 presents the major aquifer systems of the country and it is evident that the thick sedimentary successions in the Bengal Basin form prolific multi-layer aquifer systems. Thickness and lateral continuity of the different aquifers vary significantly from zone to zone, i.e. the Plio-Pleistocene aquifer occurs at a depth of around 300 m in the coastal area, whereas the same aquifer occurs only at 10 to 30 m in the Pleistocene Terraces.

2

OPTIONS FOR ARSENIC SAFE WATER

Various options are being suggested as sources of arsenic safe water involving use of rainwater, surface water and groundwater (BRAC 2000, Hoque et al. 2000, van Geen et al. 2002, GOB 2002, Yu et al. 2003, van Geen et al. 2003, GOB 2003, BRAC 2003, BRAC & WB 2003). The options can be broadly classified under two groups viz. existing sources and new sources. Different sources are briefly discussed in the following sections. 2.1 Use of existing sources 2.1.1 Well switching It has come out from arsenic test results in various areas of Bangladesh that there is enormous variability in spatial distribution of arsenic from district to village scales. In some areas, like in the districts of Chandpur, Lakshmipur, Comilla, Noakhali, Faridpur, Gopalganj, Shariatpur, Munshiganj, almost all well water exceed the BDWS whereas over most of the country there is an intimate association of safe and adjacent unsafe wells. In such areas one good option is to switch the source of drinking water. It has been reported from one area where 52% wells exceed the BDWS that almost 90% of the inhabitants live within 100 m of a safe well (van Geen et al. 2002). In such settings public awareness and dissemination campaigns can motivate the people to share the good wells with their neighbors. Monitoring becomes a crucial concern in this case, as it has been suggested that arsenic concentrations should be expected to rise in the future, even at wells that are currently safe (Burgess et al. 2002b). 2.2 Introduction of new sources 2.2.1 Surface water sources People in Bangladesh used to drink surface water from rivers, ponds, canals etc. Groundwater was introduced as drinking water in early 70s as thousands of people used to die every year due to 287

Figure 4. (a) Major hydrogeological zones of Bangladesh; and (b) Schematic NS geological cross section across western Bangladesh. Aquifers of the north: (a) Shallow Holocene fan deposits and (b) Deeper PlioPleistocene fluvial deposits (both the aquifers are arsenic safe). Aquifers of the south: (1) Upper shallow aquifer composed of Holocene fine to very fine sands (severely arsenic contaminated); (2) Intermediate Holocene medium to coarse sand aquifer with occasional gravels (sparsely arsenic contaminated); (3) Lower shallow Holocene fine to medium sand aquifer (mostly brackish water) and (4) Deep Plio-Pleistocene aquifer (fresh, arsenic safe) (after Ahmed 2003b).

water born diseases. As the tube well technology became very popular surface water sources have been abandoned. At the same time safe surface water has also become a matter of concern due to indiscriminate disposal of industrial and municipal wastes and overall poor sanitation condition in the country. Under current conditions surface water is not consumable without treatment. As surface water has been found mostly safe from arsenic, there are strong campaigns for use of surface water as drinking water. Large-scale surface water treatment plants are operational in a number of cities including Dhaka. Pond sand filters (PSF) are being considered as a source of safe water and are being installed in various parts of the country (GOB 2003, BRAC 2003). 288

Table 1.

Aquifer systems of Bangladesh.

UNDP 1982

BGS & DPHE 2001

Aggarwal et al. 2001

Composite aquifer

Upper shallow aquifer

1st aquifer

JICA 2002

GOB 2002

Arsenic status

Shallow aquifer (1st aquifer)

Upper Holocene aquifer

Shallowest part uncontaminated; lower part contaminated Most severely contaminated, peak arsenic concentrations occur here

Middle Holocene aquifer

Main aquifer Deep aquifer

Lower shallow aquifer Deep aquifer

2nd aquifer 3rd aquifer

Middle aquifer (2nd aquifer) Deep aquifer (3rd aquifer)

Lower Holocene aquifer Plio-Pleistocene aquifer

Least contaminated Uncontaminated

2.2.2 Rainwater harvesting Harvesting of rainwater is also considered as a source of arsenic safe water in Bangladesh, at least during the monsoon months as there is abundant rainfall during this time (Ahmed 2003, GOB 2003, BRAC 2003). 2.2.3 Very shallow groundwater Abstraction of groundwater in Bangladesh started with dug wells although they became almost extinct due to the overwhelming increase in the number of tube wells over the last three decades. It has been found in many places that dug well water contains arsenic at very low concentrations even in the severely arsenic affected areas (BGS & DPHE 2001). The use of dug wells as a source of arsenic safe water is being considered and various noble designs have been suggested to make them safe (Ahmed 2003, GOB 2002, GOB 2003). 2.2.4 Deep groundwater Use of deep safe groundwater is considered as one of the main options for As safe water supply (Ahmed 2003, van Geen et al. 2003, Yu et al. 2003). Deeper groundwater is As safe all over the country though the safe depth varies considerably from place to place, even at village scale (BGS & DPHE 2001, JICA 2002, van Geen et al. 2003). Deep groundwater is the most popular safe water option among the community (BRAC 2003; BRAC & WB 2003). It has been reported that one deep well, if installed strategically, can serve a population of 500 within a catchment of 300 m radius (van Geen et al. 2003). Yu et al. (2003) conclude that if the 31% of the existing wells exceeding the BDWS is replaced by deep wells, health effects related to drinking As rich water can be reduced by 70%. 2.3

Relative risks of various sources

While introducing any new source of water, the relative risk of the options has to be considered. Moreover, if not careful enough, a new risk might be substituted inadvertently while introducing a new source – risk of As in Bangladesh drinking water is a classical example. While reducing the risk of microbiological contaminants, As risk has been substituted inadvertently (Smith et al. 2000). Due to this now there are strong campaigns by some groups to return to surface water. However, there are suggestions that the shifting from tube wells should be as limited as possible by using the safe wells (Calddwell et al. 2003). The same study also concludes that the most urgent need is not changing source of water but comprehensive national water testing providing essential 289

information to households about the safe and unsafe wells. Moreover, the potential exposure to As sources other than drinking water make risk assessment of arsenic in drinking water difficult (Buchet & Lison 2000). Therefore, one has to be extremely careful in recommending alternative source of drinking water without assessing the relative risks and possible exposure through other sources. All the different sources considered for As safe water has different risks associated. It is not possible at this stage to rank different options in the absence of various quantitative matrixes. Therefore, the qualitative risk factors associated with various options are discussed here. Also findings from different studies with various options are considered in assessing the risks (Sutherland et al. 2001, van Geen et al. 2002, van Geen et al. 2003, Burgess et al. 2002a, b, Cutbert et al. 2002, Yu et al. 2003, BRAC 2003, BRAC & WB 2003, GOB 2003)

• • •

• •

•

Well switching may give rise to social problems in sharing individually owned wells; larger abstraction rate may influence the water quality; may not be suitably located for sharing by the neighbors. Removal efficiencies influenced by raw water chemistry; sludge disposal is a major issue – can create other sources of exposures; household level management more difficult; not liked much by the community. Surface water sources have high risk of microbiological contamination; pond sand filter may not remove all microorganisms; source protection is a major issue; not equally available in space and time; can have other heavy metals from industrial wastes; reports of occurrences of toxins derived from cyanobacteria; chemical and microbiological quality can vary abruptly in response to flow conditions and land use practices; not considered as the best option by the community. Rainwater is unequally distributed in time; storage is a major issue; quality deterioration during long-term storage is a matter of concern; long-term consumption can lead to element deficiency; not liked by the community. Dug wells can not be developed under all types of geological and hydrogeological conditions; there are risks of microbiological contamination if not well protected and placed at safe distances from existing sources of pollutions; there are reports of As occurrences above the BDWS; other chemical parameters such as nitrate and manganese can also be limiting factors; can become dry during drought/low water table conditions; not considered as a best option by the community. Deep tube wells can have arsenic and other chemical contaminants in some areas; can induce leakage of As rich water from upper aquifers if pumping rate is high; poor construction can result into short circuiting of arsenic rich water; arsenic may be released from aquifer sediments under changed hydrogeochemical conditions imposed by new pumping regimes; possibilities of resource depletion if abstraction is not regulated.

It is evident from the discussions above that none of the option is risk free. In selecting a particular option one needs to consider the relative microbiological and chemical quality factors along side sustainability, affordability and acceptability by the community. Community options rather than household options should get priority for the ease of management and monitoring. Use of deep or safe community wells seems to be the most practical option available to reduce the arsenic exposure, also this is most liked by the community (van Geen et al. 2003, Yu et al. 2003, BRAC & WB 2003). If piped water systems are introduced in smaller urban centers and rural areas, as outlined in the arsenic mitigation policy of the government, the same source can be used.

3

MANAGEMENT OF ARSENIC ENRICHED GROUNDWATER

The issue of arsenic management is very wide and includes management of water resources, sources of water supplies, provision of alternative sources, health aspects, social aspects, agricultural issues etc. Though all these are important and one can not speak of arsenic management without a holistic approach, many of the issues are outside the subject matter of the paper. And, therefore, the issue of arsenic enriched groundwater management is highlighted here. 290

3.1

Arsenic in existing national policies

In Bangladesh there are number of modern policies which can be utilized in management of As enrichment of groundwater. The policies adopted prior to detection of arsenic naturally do not address the issue. However, the recently adopted policies address the issue and the most relevant ones are the National Policy for Safe Water Supply & Sanitation 1998 and National Water Policy 1999. These two policies lay the foundation of groundwater management in the country. A new policy focusing only of arsenic, National Arsenic Mitigation Policy 2004, has been adopted recently by the council of ministers. Also National Environmental Policy and National Environment Management Action Plan 1992 can be used to introduce relevant legal instruments for arsenic management. National Policy for Safe Water Supply & Sanitation 1998: The policy sets a goal for the government to ensure that all people have access to safe water and sanitation services at affordable cost. In the policy document safe water supply is defined as ‘means of withdrawal or abstraction of either ground or surface water as well as harvesting rainwater water, its subsequent treatment, storage, transmission and distribution for domestic use”. The goals and objectives includes, among various other issues, reduction of incidence of water borne disease, ensuring supply of quality water through observance of accepted quality standards and removal of arsenic from drinking water and supply of arsenic free water from alternative sources in arsenic affected areas. The strategy of the policy is set to development goals based on a number of principles including precedence on safe water from surface sources. The strategy also emphasizes on regular and coordinated water quality surveillance to controlling and preventing contamination of drinking water. The policy asks for excavation or re-excavation and preservation of at least one pond in each and every villages of Bangladesh with necessary security measures to prevent water of the pond from contamination. For urban water supplies it also emphasizes on monitoring of water quality for the purpose of ensuring an acceptable standard. National Water Policy 1999: The national water policy is declared “to ensure continued progress towards fulfilling the national goals of economic development, poverty alleviation, food security, public health and safety, decent standard of living for the people and protection of natural environment”. The policy will guide management of water resources of the country by all concerned agencies. The policy states “the ownership of water does not vest in an individual but in the state. The Government reserves the right to allocate water to ensure equitable distribution, efficient development and use, and to address poverty”. Under the water supply and sanitation section the policy identify number of problems related to groundwater, viz. arsenic contamination, heavy withdrawals for irrigation and subsequent decline in water table, seepage of agrochemicals into shallow aquifers, and salinity intrusion. On the other hand, in the water and agriculture section, it encourages future groundwater development for irrigation subject to regulations prescribed by the government from time to time and without affecting drinking water supplies. It also emphasizes strengthening monitoring organizations for tracking groundwater recharge, surface and groundwater use, and changes in surface and groundwater quality. The policy talks about research, central database and enactment of a national water code revising and consolidating the laws governing ownership, development, appropriation, utilization, conservation, and protection of water resources. 3.2

Institutions involved

In Bangladesh there are a large number of organisations involved in water resources development and management. However, currently most of these organisations develop water only, very little management is done. The National Water Policy outlines the institutional arrangements for future integrated water resources management in the country. The Ministry of Water Resources is the focal point in the public sector for water resources management. However, the arsenic issue is currently considered mainly as a water supply issue and therefore, the Ministry of Local Government has been made the focal ministry. The other ministries involved are Ministry of Health and Family Welfare, Ministry of Water Resources, Ministry of Agriculture, and Ministry of Science and Information Technology. There is a high power secretarial committee to co-ordinate the activities 291

Table 2.

Major government organizations involved in arsenic enriched groundwater management.

Sl no. Organization

Main role

Role in arsenic

1

Department of Public Health Providing water supply and Engineering (DPHE) sanitation all over the country except two major cities

Testing water sources, finding new sources, providing new sources

2

Bangladesh Arsenic Mitigation Water Supply Project (BAMWSP)

Specifically raised to deal with arsenic mitigation

Testing wells, disseminating results, providing arsenic safe water

3

Bangladesh Water Development Board (BWDB) Bangladesh Atomic Energy Commission (BAEC)

Monitoring of water resources, Investigation of arsenic source development of surface water, in sediments and to some extent ground water Use of nuclear technology Use of isotope hydrology in managing arsenic enriched groundwater

5

Geological Survey of Bangladesh (GSB)

Preparing geological maps, conducting geological exploration

Aquifer mapping

6

Bangladesh Council for Scientific and Industrial Research (BCSIR)

Conducting applied research for developing new technologies

Validation of arsenic test kits and removal technologies

7

Water Resources Planning Organization (WARPO)

Preparing water resources master plan for the country

Integration of issue of arsenic in overall water resources management

8

Bangladesh Agricultural Development Corporation (BADC)

Providing support in the field of agriculture

Assessment of water resources and impact of arsenic on irrigated crops

9

Bangladesh Institute of Nuclear Agriculture (BINA)

Application of nuclear technology in agriculture

Impact of arsenic in food chain

10

Bangladesh Rice Research Institute (BRRI)

Development of rice

Impact of arsenic in food chain

4

of different ministries in this field. Apart from government organizations (Table 2), various research organizations, non-government organizations, development partners, international organizations are also involved. The involvement of so many stakeholders in the issue makes it very complicated and co-ordinated efforts may become difficult. 3.3

Mitigation strategy

Though there are various options of safe water available, the mitigation efforts implemented so far is insignificant compared to the magnitude of the problem. This, to some extent, is due to lack of a mitigation strategy. There are various suggestions regarding the mitigation strategy (GOB 2002, Nuruzzaman & Ahmed 2003, Alaerts & Khouri 2004). All these suggest various approaches for arsenic mitigation encompassing water supply, health, social and food chain issues. However, the recently adopted National Arsenic Mitigation Policy will be considered as the main instrument for managing the big problem. The main policy statements included in the policy are:

• • •

access to arsenic-safe water for drinking and cooking will be ensured; all patients will be managed effectively; public awareness will be raised about impact of arsenic contaminated water; 292

• •

capacity will be built at all levels for implementation of mitigation options, surveillance and monitoring of water quality and diagnosis and management of patients; impact of arsenic on agriculture will be assessed.

Although the policy gives priority to all the major issues related to arsenic it could be made more pragmatic. The issues which should are important in setting long term goal for arsenic management are discussed in the following sections. 3.4

Requirements for sustainable management

3.4.1 Concept of safe drinking water In the arsenic mitigation policy emphasis has been given on arsenic safe water but one should introduce the safe water concept. In Bangladesh drinking water safety is often determined based on single or few parameters, e.g. in the past microbiological contaminants were the most focused parameter. It is unrealistic for any water supply system to consider one or two parameters. All the health sensitive parameters should be considered while introducing a new source of drinking water and at the same time all existing sources should be checked for all parameters. 3.4.2 Reliable testing facilities One major issue in the field of drinking water arsenic occurrence is reliable detection at the desired levels. Hundreds of thousands dollars have been spent in testing all existing water sources in the arsenic prone areas by field kits. There are strong reservations about the reliability of such kits. However, recent experiences show that the kits are good in delineating the overall pattern of occurrences. Also the kits can detect wells exceeding the BDWS. It is important to have reliable testing facilities available a village levels. 3.4.3 Legal aspects Although there are number of policies regarding water management, safe drinking water and arsenic management, there are no legal bindings for the water supply authorities to ensure quality of supply. It is important to introduce a “Safe Drinking Water Act” under which health-based standards can be set to protect drinking water from naturally occurring and anthropogenic contaminants. This act, like the US SDWA (EPA 1999), can ensure the right to know what is in the drinking water. Also this can make provisions so that all water suppliers notify consumers quickly when there is any problem with the water quality. 3.4.4 National monitoring Although it is mentioned in the major national policies, the water supply systems in Bangladesh lack a systematic national surveillance and quality monitoring. It is argued that if there has been such a monitoring, the arsenic problem might have abated much earlier (Ahmed & Ravenscroft, 2000). There is a national monitoring system currently run by BWDB, which is too small compared to the number of drinking water sources. BADC has also some monitoring focused more on agricultural uses. It is therefore very important to immediately develop a national water quality monitoring and surveillance under DPHE. Monitoring systems of BWDB and BADC can feed in additional information to the proposed monitoring. 3.4.5 Research and training There are number of gaps in the understanding of the arsenic problem and needs advanced research. More applied research is needed to fill in the knowledge gap. Overall, there is shortfall of trained personnel for water resources management. Training should be given to stakeholders in all levels for efficient management of drinking water sources in particular. 3.4.6 Local capacity building Personnel working in relevant government organizations and NGOs need basic hydrogeological training. Strengthening of university departments and training organizations is an important 293

aspect of local capacity building. Not only this, there are lack of local capacity in many other areas of water resources management. Enhancement of local capacity is a key issue for sustainable management of water resources. 3.4.7 Access to information Access to information is a fundamental right to water users. However, it is very difficult under current system. BAMWSP has a arsenic management data base center to cater with all information needs in the field of testing and mitigation. However, the access is very limited and should be expanded at village levels. 3.4.8 Public awareness The current level of awareness regarding arsenic in groundwater is generally not very high (BRAC 2003). However, it is also found that people respond quickly as the information is disseminated and new sources are introduced (van Geen et al. 2003). Overall, people perception about water need to be changed. Water, in most cases, is not considered as an important natural resource and a commodity. Media can play a major role in creating awareness about quality and quantity of water resources. 3.4.9 Role of private sector Currently about 90% of the water sources are privately installed and owned. It is therefore, very important to involve the private sector in all water management issues. 3.4.10 Organizational reform Currently there are many organizations involved in the field of water resources development. However, there is no organization solely responsible for management of groundwater resources of the country. Groundwater is very important in Bangladesh and socio-economic developments are very much driven by the use of the hidden resource. It is very important to create an organization to manage the vital resource.

4

CONCLUSIONS

Management of arsenic rich groundwater is a big challenge for Bangladesh. The magnitude of the problem is very big where about one third of the existing wells exceed the current BDWS. Groundwater has been playing major role in providing access to safe drinking water and ensuring food security since the emergence of Bangladesh. There are currently moves to stop use of this strategic resource. It is not realistic and groundwater should play a vital role again in providing alternative sources of safe water. Proper groundwater management is needed along with national water quality monitoring and surveillance. Holistic approach should be taken to manage the arsenic issue. This would need major steps to be taken in the field of local capacity building, organizational reform and legal instrumentations. Concepts of safe drinking water rather than arsenic safe drinking water have to be adopted. While introducing a new source one needs to take every possible measures not to substitute a new risk. Public awareness about the quantity and quality of groundwater should be raised. Citizen’s right to access to information about water quality has to be ensured. Arsenic mitigation should be made part of the overall water resources management.

ACKNOWLEDGEMENTS I gratefully acknowledge the help of M. Ashraf Ali Seddique and M. Abdul Hoque for their help with the figures. Thanks also due to Dr. Willy Burgess for his review comments. 294

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Strengthening water examination system in Bangladesh H. Jigami Japan International Cooperation Agency (JICA) Expert, Department of Public Health Engineering, Dhaka, Bangladesh

ABSTRACT: Screening test has been implemented in order to survey arsenic contamination in Bangladesh since the first arsenic detection in groundwater occurred in 1993. National Policy for Arsenic Mitigation 2003 was approved in March 2004, and emergency response is expected to start in the country. Japan International Cooperation Agency (JICA) contributes for Department of Public Health Engineering in the field of arsenic mitigation including water examination system because accurate analysis is expected instead of field test kit which is used for screening test. Asia Arsenic Network surveyed groundwater at tube wells in 2003 as well as its first round in 2002. Fifty two percent out of 1197 green painted tube wells turned into red where arsenic concentration exceeded 50 ␮g/L. This report illustrates that resurveyed results at sites caution unstable situation in groundwater and necessity of monitoring. JICA and JICA expert’s activities are also discussed.

1

INTRODUCTION

In Bangladesh the presence of arsenic in groundwater was first detected in 1993. The estimated population exposed to arsenic contamination exceeding 50 ␮g/L in Bangladesh is 29 million (Armed, 2002). This is approximately 25% of the total population. Many international organizations and individual donor agencies are contributing to solve this arsenic crisis in many fields. Department of Public Health Engineering (DPHE) is mandated to provide safe drinking water and sanitation facilities for all over the country except for Dhaka City and Chittagong City. This paper discusses: (1) Current situation of screening tests, (2) Japan International Cooperation Agency’s (JICA) contribution to arsenic mitigation program, (3) Asia Arsenic Network’s (AAN) facts and caution from sites and (4) JICA expert’s research project for safe water.

2 2.1

SCREENING TEST AND NATIONAL POLICY Outline of screening

Bangladesh Arsenic Mitigation Water Supply Project (BAMWSP) implemented screening tests of tube wells in 1999 (Phase I) and in 2000 (Phase II) in order to understand the extent of arsenic contamination of tube wells and to identify arsenic affected people having visible signs of arsenicosis (BAMWSP, 2003). BAMWSP project was conceptualized with the joint effort of the Government of Bangladesh (GOB) and The World Bank – Swiss Agency for Development & Co-operation (WB-SDC) (BAMWSP, 2003). Phase I illustrated that 48.2% of tube wells (out of 80,368) in six Upazilas (sub-districts) were contaminated by arsenic in excess of national standard (50 ␮g/L), and 1139 people (0.07%, out of 1,698,410) were confirmed as arsenicosis patients. Phase II, in collaboration with DPHE, showed that arsenic was detected from tube well water in 45.4% (out of 589,389 tube wells) in 36 Upazilas, and arsenicosis patients were counted as 9701 (0.1%, total population in this area is 9,368,533). In Bangladesh there are 507 Upazilas (Bangladesh Bureau of 297

Table 1.

Summary of emergency and medium-term response.

Selection of villages Mitigation approach Service level Cost sharing Site selection Institution arrangement

Emergency response

Medium-term response

⬎80% contaminated wells Supply driven 50 families/water source Capital: No cost share O&M: borne by the users Discussion between supply agency & community By projects and donors accommodation, or by DPHE

80% ⬎ contaminated wells ⬎ 40% Demand driven 25 to 30 families/water source Capital: by affordability O&M: borne by the users Community’s decision By local government institution

Statistics, 2002). BAMWSP established information system as National Arsenic Mitigation Information Centre (NAMIC). The Centre was aimed to be the bank of all arsenic related information (NAMIC, 2004). 2.2

National Policy for Arsenic Mitigation

The Government of Bangladesh is preparing “National Policy for Arsenic Mitigation 2003” in order to plan and implement practical projects for arsenic mitigation (National Policy, 2003). This policy was approved by the cabinet in March 2004. This policy defines three levels of contamination magnitude in each Upazila (ward or village), such as emergency, medium-term and long-term (Table 1). 2.2.1 Emergency response The government shall focus on ensuring at least one safe source of drinking water within a responsible distance on an emergency basis.

• • • • • • •

Selection of villages: More than 80% of groundwater from tube wells is contaminated. Mitigation approach: Emergency response shall be based on a supply driven. Community and local government institutes will be engaged and take responsibility of Operation & Maintenance. Service level: Emergency response shall ensure safe water source, and will deal with only community facilities. Cost sharing: There shall be no cost sharing for the capital cost of facilities. Operation & Maintenance cost shall be borne by the community. Site selection: The supply agency should discuss the location of the safe source with the community. The location should be agreed considering gender, patients and others. Institutional arrangement: Projects and (donor) agencies shall be required to accommodate mitigation approach. If any projects and/or (donor) agencies are not available, DPHE shall implement the emergency program in the area. Audit: National level research and development organisations shall audit in order to verify services.

2.2.2 Medium-term response The medium-term arsenic mitigation program shall be completed within a period of three years.

• • •

Selection of area: More than 40% and less than 80% of groundwater from tube wells are contaminated. Mitigation approach: Medium-term response shall be based on a demand basis. It implies active involvement of the community and devolution of responsibility to the Local Government Institutes. Cost sharing: Capital cost shall consider the affordability. Total cost of Operation & Maintenance shall be borne by the users. 298

• • •

Service level: Community facility shall be provided at the rate of one water point for a cluster of 25 to 30 households. Site selection: The community that shall use the facility shall select the site for the installation of the facility. The site shall be such that it facilitates the access of the poor. Institutional arrangement: The local government institutions, particularly the Union Parishad, shall play more effective role in the medium-term water supply program.

2.2.3 Long-term response The long-term response to arsenic mitigation shall consider the same criteria as under Mid-Term response. The difference being that in this phase arsenic mitigation programs shall promote proven and sustainable technology options. This phase shall include many of the technology options promoted during the mid-term response. The activities towards long term response shall be initiated immediately. The long-term response shall also include piped water supply systems for the rural areas. 2.3

Technology options

The “National Policy for Arsenic Mitigation 2003 (Draft)” explains also technology options for safe water supply. Arsenic mitigation programs for safe water supply shall promote arrange of options. The implementation plan describes these six safe water supply options (National Policy, 2003). 2.3.1 Improved dug well The first option for water supply is Improved Dug Well. Sand Filter shall be provided in case of turbidity. “Guideline for Construction of Dug Well” is provided by expert committee in Bangladesh. This option cannot be used in areas with loose sandy soil, salinity groundwater and stony hills. Although the Policy does not mention this aspect, bacteria removal shall be also expected by this sand filter. 2.3.2 Surface water treatment 2.3.2.1 Pond sand filter Treatment of natural pond water with slow sand filtration shall be considered a preferable option. This is popularly known as Pond Sand Filter (PSF). PSF is low cost and efficiently removes turbidity and bacteria. The expert committee provides guideline as “Guidelines for Construction of PSF.” Arsenic removal is not a main purpose of PSF but an example shows that arsenic concentration of 16 ␮g/L in raw water is reduced by 1 ␮g/L in filtered water. 2.3.2.2 Larger scale surface water treatment Surface water treatment from flowing rivers and large water bodies by appropriate treatment plants shall be encouraged for large-scale water supply. 2.3.3 Deep hand tube wells When conditions are feasible, Deep Hand Tube Wells could be tried. Conditions shall be (a) impermeable layer separates the deeper aquifer from the contaminated upper aquifer and (b) Pond Sand Filter and Dug Wells are found to be technologically not feasible. Deep Hand Tube Wells should follow “Protocol for Sinking Deep Tubewells in Arsenic Contaminated Areas.” 2.3.4 Rainwater harvesting Rainwater Harvesting has good potential especially in salinity affected areas. In the context of cost and limitation of long term storage of rainwater, this approach aims at persuading people to drink rainwater at least in rainy season. Being a household based option it is recommended that the government’s role in Rainwater Harvesting should be limited to promotional activities. 299

2.3.5 Arsenic removal technology Bangladesh Council for Scientific and Industrial Research (BCSIR) verified four technologies as arsenic removal technologies in February 2004 as the first phase. After the validation of a given technology it is for the private sector to develop marketing and distribution networks. The role of the government shall be limited to validation of the options and making provision of legal instruments for consumer safeguards against all adverse affect resulting from the use of the options. National Committee of Experts on Arsenic shall develop detailed protocols regarding sludge and spent media removal. Continuous monitoring of such removal options by the public sector agencies shall be ensured. 2.3.6 Piped water supply system The long-term goal should be to introduce piped water supply systems both in the rural and urban areas preferably based on surface water treatment plants. For the rural areas government shall facilitate testing and piloting of small-scale piped water supply systems based on improved dug well, Pond Sand Filter or other surface water and safe groundwater sources. Since the National Policy has been approved finally, it is necessary to decide how to implement these emergency response and other actions. Unfortunately it is deduced that the Government of Bangladesh is not able to budget for all responses in short-terms. Funds from outside should be crucial for quick actions. JICA will discuss with “Unit for Policy Implementation” (UPI, Danida) and “Arsenic Policy Support Unit” (APSU, the government of Bangladesh and DFID) as donor cooperation. 2.4

Field test kits

Screening tests adopted Field Test Kits because of its convenience, such as quick and easy procedures. Disadvantages of field test kits are listed up as follows:

• •

Human errors: After adding reagent to water sample, operator or examiner judges its colour visually. If judging standard or rule is not completed to every operator, human errors would always occur. Accuracy: Reaction of reagent shows its result by colour. Operator or examiner is required to judge and select the figure of arsenic concentration among followings.

BDL (Below Detected Level), 0.01, 0.02, 0.05, 0.1, 0.2, 0.3, 0.5 and 0.7 (NIPSOM). Example: When colour of sample shows between 0.05 and 0.1, operator should choose only one result. This means that result of 0.05 would be in range between 0.035 and 0.075. This kind of human errors will never be avoided unless water quality analysis adopts a more quantitative method instead of field test kits which use colorimetric judgement.

3

JICA’S COLLABORATION WITH DPHE

3.1 Emphasis in arsenic mitigation JICA understands that arsenic mitigation is one of the most important issues in Bangladesh in terms of bilateral cooperation. Especially accurate water quality analysis should be crucial after screening tests of tube wells instead of field test kits. When donor agencies and DPHE install water supply options in communities or villages, it is required to ensure the water quality of new options. Therefore JICA lays emphasis on accurate water quality analysis. DPHE requested a grant aid project to the Government of Japan in 2002. JICA is collaborating with DPHE in “Strengthening Water Examination System of Bangladesh” for DPHE. DPHE has four existing regional laboratories for water quality analysis in Comilla, Khulna, Mymenshingh and Rajshahi, and has analysis rooms in Jhenaidah (assisted by JICA) and Noakhali (by Danida). BAMWSP supported (1) interior renovation of these laboratories and (2) installation of Atomic Absorption Spectrophotometer (AAS). New DPHE laboratories will be established by BAMWSP in 300

2004 in Barisal, Rangpur and Sylhet, and another two laboratories are coming in the location of Bogra and Tongi. Unfortunately there is no central laboratory of DPHE for coordination and management. 3.2

Concept of Central Laboratory

Japan Grant Aid Project “Strengthening Water Examination system of Bangladesh” consists of (1) establishment of Central Laboratory and (2) upgrade of existing analysis rooms in Jhenaidah and Noakhali. In this section, discussion is focused on central laboratory. Main roles of Central Laboratory will be as follows: 1. 2. 3. 4. 5. 6.

Management and coordination of regional laboratories QA (Quality Assurance)/QC (Quality Control) of water quality analysis Analysis of special parameters which regional laboratories cannot analyse Training of chemists and analysts Maintenance and support of equipment and consumable items Data Management (Establishment of Database).

Tentative schedule illustrates that Basic Design Study will end in July, 2004, Detail Design Study will start in November 2004 after the cabinet approves grant aid for construction, and Construction of Facility will start in 2005 and end in 2006. Japan charges mainly in this Grant Aid project (1) design of facilities, (2) construction of facilities, (3) supervise of construction, (4) procurement of equipment, and (5) consultation of Central Laboratory including donors’ coordination. On the other hand DPHE pledges that (6) setup of new division for water quality analysis within DPHE structure, (7) recruitment of personnel for Central Laboratory under revenue budget and (8) cost of operation and maintenance. For sustainable laboratory work, DPHE is preparing the revolving system of water quality analysis. This is the most welcome action for the local government in terms of sustainability. When this revolving system is approved by the government, each laboratory could keep certain percentage of analysis fee with them and utilise this budget for operation and maintenance purposes. When DPHE needs assistance for technical and management support to be a going concern, JICA will consider this type of project for recipient organisation.

4

CAUTION FROM SITE

JICA has a project partner, Asia Arsenic Network (AAN), in Bangladesh. AAN is Japan based NGO and has a lot of experience and expertise in awareness and arsenic mitigation activities. AAN participated in screening test in Jessore District in western Bangladesh, and examined tube well water using NIPSOM Field Test Kit in 2002. One of unique activities of AAN is “Mobile Arsenic Centre,” and it consists of: (a) confirmation of arsenicosis patients by medical doctors; (b) promotion of arsenic poisoning awareness by music and short plays, (c) presentation of arsenic poisoning cases in other countries, and (d) introducing nutrition advice for prevention of arsenicosis. In “Mobile Arsenic Centre,” chemist of AAN examined the same tube well water in second round although other agencies and NGOs do not resurvey the same tube wells. 4.1

Screening test as first round

In 2002 from February to August, the first round screening test was implemented in Sharsha Upazila in Jessore District. Sharsha Upazila consists of 11 unions, and each union has nine wards. AAN adopts “NIPSOM” Field Test Kit for screening test. Reagents such as potassium iodide, tin chloride and zinc are added in 10 mL of water sample from tube well. After five minutes, analyzer judges concentration with sample’s colour in comparison with standard colour chart. AAN measured arsenic concentration at 32,441 tube wells in this Upazila. The number of tube wells which showed 50 ␮g/L and below was 24,879 (76.6%). These tube wells are painted in 301

Table 2.

Screening test as first round.

Name of union 1 2 3 4 5 6 7 8 9 10 11

⬍50 ␮g/L

⬎50 ␮g/L

Rate of contamination (%)

2494 1860 2301 2469 1937 2537 1993 2959 3133 4914 5844

1376 768 1755 2257 1632 2249 1411 1371 2.757 3936 5367

1118 1092 546 212 305 288 582 1588 376 978 477

45 59 24 9 16 11 29 54 12 20 8

32,441

24,879

7562

23

No. of tested wells

Kayba Goga Putkhali Bahadurpur Lakshmanpur Dihi Nizampur Bagachra Ulashi Sharsha Benapole Total Table 3.

Results of arsenic concentration in second round. No. of tube wells resurveyed in 2003 (␮g/L)

Results in 2002 (␮g/L)

BDL

10

20

50

100

200

300

400

500

BDL 10 20 50

0 3 0 0

22 136 91 24

8 105 141 106

0 36 71 69

0 17 37 29

1 13 54 63

0 14 40 48

0 8 16 29

0 2 7 7

Table 4.

Results of second round.

Resurveyed wells

Group 1 (same)

Group 2 (decreased)

Group 3 (Increased)

1197

346

224

627

green. Tube wells which is exceeding 50 ␮g/L were 7562 (23.3%), and are painted in red. Table 2 shows details of screening test as first round in Sharsha Upazila. AAN set up a concept for re-survey as follows (Uddin et al., 2003): Select wards in which more than half of tube wells show high arsenic contamination in excess of 50 ␮g/L. The number is 18 wards. In these 18 wards, tube wells should have showed concentration as 50 ␮g/L and below in 2002. These tube wells were painted in green. 4.2

Resurvey of groundwater

Table 3 illustrates arsenic concentration in groundwater in 2003 as second round. The first column shows arsenic concentration in 2002 (first round). In the same row, the result of second round is shown. For example, shadowed cells mean there is no change of arsenic concentration in this year. According to the results, tube wells are categorised into three groups: (1) Group 1: The same arsenic concentration, (2) Group 2: Decreased arsenic concentration, and (3) Group 3: Increased arsenic concentration. Table 4 summarises resurveyed tube wells and its arsenic concentration. 4.3

Conclusion

It can be concluded from resurvey of groundwater as follows. (1) Half (627 out of 1197) of green coloured tube wells turned into red coloured ones in next year. (2) Approximately 19% (224 out of 302

1197) of tube wells decreased arsenic concentration and all of them were below national standard and (3) 29% of tube wells showed the same range of arsenic concentration as the previous year. It is deduced that arsenic concentration in groundwater is not stable for long time. Even if groundwater quality is safe once, it might not be safe next year. Although the mechanisms of arsenic leaching is not clear, arsenic concentration may be affected by changing level of water table, migration of groundwater, groundwater usage for irrigation and other natural and human factors. Therefore, continuous observation as monitoring of groundwater quality should be crucial, and groundwater quality analysis should be accurate in some occasion. It is highly recommended that water quality analysis be operated periodically and accurately.

5

RESEARCH FOR APPROPRIATE TECHNOLOGY

5.1

Background of research

In terms of safe water supply options in Bangladesh, several options are introduced and implemented already in some areas. As mentioned earlier, each technology option seems to meet every local condition but there is no perfect technology option. It may happen that community or village has no choice of technology option of safe water supply. If there is area where any option can be adopted, there seems to be no possibility of water supply to this area. Some villages in Jessore District have the following conditions: Example: Loose sandy soil: not suitable for dug wells Saline water: not feasible for deep hand tube well Low precipitation: no potential for rainwater harvesting This area has (1) many natural ponds and (2) highly arsenic contaminated tube well water. Some facts are confirmed in sites and in other researches through discussion with AAN.

• • •

Although groundwater for irrigation contains arsenic, soil at paddy fields does not accumulate arsenic as predicted by the theory (Morikawa et al., 2002). Generally, water in natural ponds does not contain so high arsenic concentration. Research papers report that certain types of micro organisms decompose arsenic compounds into gaseous substances (Hiroki, 1996).

5.2

Hypotheses

According to these facts, the hypotheses proposed by the JICA experts are as follows:

• •

Certain types of micro organisms decrease arsenic concentration of groundwater at natural ponds. People can use safe drinking water after natural treatment at ponds through filtration.

5.3

Outcomes

JICA expert started research project in collaboration with Bangladesh Engineering & Technological Services (BETS), University of Dhaka and AAN. AAN has kindly arranged natural ponds for this research. After approximately 30 days of observation, these results can be confirmed through laboratory and field experiments.

• •

In laboratory, when arsenic contaminated water contacts with soil which contains numberless micro organisms, concentration of arsenic decreases. In field, after arsenic contained groundwater is transferred to natural ponds, arsenic concentration decreases. 303

•

In laboratory, when micro organisms were isolated from soils, many types of micro organisms decreased arsenic concentration.

Although this research was constricted in its time, it is clearly confirmed that certain type of micro organisms decreased arsenic concentration in the circumstances of natural ponds. 5.4 Next steps It is deduced that application of micro organisms has high potential of appropriate technology, but there are still many subjects that should be confirmed:

• • • 6

Identification of micro organisms Side effects of by-products or decomposed substances, and Confirmation of appropriate conditions CONCLUSIONS AND RECOMMENDATIONS

6.1 Conclusions This paper discusses National Policy for Arsenic Mitigation in Bangladesh, expected alternative water options, examples of arsenic concentration in groundwater, Government of Bangladesh’s project in collaboration with JICA, and research work using micro organisms. It is concluded as follows:

• • • • • •

Water quality of groundwater is not stable. Accuracy of field test kits is limited although they are useful for easy and quick testing. There is no system for water quality analysis at DPHE in Bangladesh except for individual laboratories. The government of Bangladesh proceeds arsenic mitigation planning and practical projects under National Policy. JICA is supporting the government efforts in mitigating arsenic including establishment of water examination system towards accurate analysis. It is confirmed that micro organisms in natural ponds can reduce arsenic concentration in groundwater through laboratory experiment and field survey.

6.2 Recommendations According to these conclusions the author strongly recommends the following points:

• • •

The government of Bangladesh and donor agencies are to be encouraged to plan mitigation projects including water quality analysis in the country. DPHE will analyse water samples in practical ways. Accurate and reliable analysis system should be established at DPHE. Field test kits are also useful for the purpose of easy and quick testing. Periodical monitoring of groundwater should be crucial especially in highly arsenic contaminated areas.

REFERENCES Armed, M.F. 2003. Alternative Water Supply Options for Arsenic Affected Areas of Bangladesh. In: Arsenic Mitigation in Bangladesh. Ministry of LGRD and Cooperatives, Government of Bangladesh. BAMWSP 2003. Status Report of Bangladesh Arsenic Mitigation Water Supply Project, Dhaka Bangladesh. Bangladesh Bureau of Statistics 2002. 2000 Statistical Yearbook of Bangladesh 21st Edition, Dhaka Bangladesh. Government of Bangladesh 2003. Draft National Policy for Arsenic Mitigation 2003, Dhaka Bangladesh. Hiroki, M. 1996. Micro organisms in arsenic contaminated soil. Soil Microorganisms 47: 23–30 (In Japanese).

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Hiroki, M. 1993. Effect of Arsenic Pollution on Soil Microbial Population, Soil Sci. Plant Nutr. 39(2): 227–235. Hiroki, M. 1993. As-Tolerant Bacillus circulans Isolated from As-Polluted Soils. Soil Sci. Plant Nutr. 39(2): 351–355. Hiroki, M. & Toshiwara, Y. 1993. Arsenic Fungi Isolated from Arsenic-Polluted Soils, Soil Sci. Plant Nutr. 39(2): 237–243. Kawahara, K. 2002. Current situation of arsenic contamination in agricultural crops. JICA Expert Report KK-JR-33, (In Japanese). Morikawa, H. et al. 2002. Arsenic Distribution of groundwater and grand in Marua Village, Bangladesh, 7th Forum on the Arsenic Contamination of Groundwater in Asia, Miyazaki, Japan. NAMIC 2004. National Arsenic Mitigation Information Centre Home Page, URL: http://www.bamwsp.org/ osman/namic.htm (Accessed on 6 March 2004). Uddin, M. &. Shamim, M. 2002. Highly Contaminated Areas Arsenic Safe Tubewells Re-Screening Survey Reports in Sharsha Upazila of Bangladesh, JICA/AAN Arssenic Mitigation Project, Bangladesh.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Implementation of safe drinking water supplies in Bangladesh C.F. Rammelt & J. Boes Faculty of Technology, Policy and Management, Delft University of Technology, Delft, Netherlands

ABSTRACT: The normal focal point with the introduction of ‘new’ technologies is the technique itself despite numerous experiences that show that most problems evolve around its implementation process. This paper uses the arsenic contaminated drinking water supply in rural Bangladesh to illustrate our analysis of this discrepancy and presents recommendations and fundamental principles for promising approaches. To value these it is important to look at a few basic perceptions of current paradigm first. Today’s paradigm supports the ideas of growth, investments in capital and technology transfer fuelled by a one-sided western view on development. Apart from the consequences of operating within this paradigm, the implementation of solutions in the case of the arsenic issue is further hampered by the numerous problematic elements such as: lack of awareness, general water scarcity, doubts about the causes of the contamination and other research uncertainties, etc. This paper will however not focus on these elements separately but three causal relations between them: (1) Physical oriented perspective: The current situation of the drinking water supplies reveals a discrepancy between the scarce availability of safe drinking water and the needs of the rural community; (2) Knowledge oriented perspective: The health hazards issue shows the relation between research experience and their effects on local coping mechanisms; (3) Institutional oriented perspective: The problem of lacking proper drinking water supply institutions relates the institutional bottlenecks with the effects on local arrangements. Starting from these different perspectives this paper tries to give a holistic basis from where recommendations and approaches can be formulated. As the technological solutions are not considered to be the main bottlenecks today, the recommended approaches particularly focus on the social and institutional level of implementation in an interdisciplinary way.

1

INTRODUCTION

Introducing some basic development conceptions of current paradigm is useful to understand the way the arsenic problem is approached (Rammelt & Boes 2002). This will be the focus of the first paragraphs. The recommendations that surface from this analysis are presented in the second half of the paper. The undertone in the ideas about development was set during the colonial era. In that time it became a widely held belief that poor countries were on their way in a development path comparable to that of the rich countries. This assumption in mainstream growth and development concepts was used to justify capital and technology transfer to the stigmatized ‘underdeveloped’ countries, because such investments were believed to be the key to growth. However, technology functions only when it is an integral part of a system with appropriate cultural and institutional conditions. The mainstream approach often falls short in this respect and eventually the failures of this top-down method of the 1970s became relatively easy to spot. However, there was another, less obvious side to consider as well. With their technological advance, industrialized nations often unwillingly contributed in strengthening dependencies and widening the gap between rich and poor. Not only on a global level but also within the ‘receiving’ country itself. In the way a technology is transferred it can create huge problems such as a 307

reinforcement of current power relations between producers and workforce, a disturbance of the local markets and contrasts between rural and urban development. In brief it can intensify the duality within a society. This does however not imply discarding Western technologies altogether in a country where they have not been developed. The application of technology to areas not necessarily related to the initial implementation should however be done with great care. Kranzberg summarized this statement by saying that technology is neither good nor bad, nor is it neutral (Kranzberg 1986). We could add that it is the way a technology is applied that sets the value on it, and that makes it appropriate or not. Just after decolonization a common attitude of the industrialized world towards their former colonies was that these simply couldn’t manage independently. But there was no form of cooperation and, derived from the legacy of the colonial rule, planning systems had a top-down conception and were something governments did for the people. With the Second World War, the language became military and bureaucratic with a jargon based on terms as ‘objectives’, ‘targets’, and ‘strategies’. The many failures of top-down tactics eventually produced a shift towards bottom-up or grassroots approaches based on partnerships with local communities in the project areas. So-called participatory development arose and was popularized particularly by amongst others Robert Chambers (Chambers 1992) and David Korten (Korten 1996). A more recent theory, Participatory Action Research works directly with local capacities to bring about changes in power relations. However, other problems might arise when outsiders interfere and try formalizing local organizations without properly understanding the local power structures. This has often occurred when the informal familial, gender, professional arrangements could not be openly shared with outsiders and the actual decisions are taken behind the scenes. Nevertheless, a key criticism to participatory approaches in general remains that groups are formed within the local community in order to meet predetermined objectives, at best they might participate in a more or less joint analysis, leading to the formation of new local institutions that will eventually take over control of local decision-making. Development conceptions must start appreciating the non-project nature of people’s lives. Therefore, in this paper we will argue that an important part of research efforts should focus on the effects of technological and institutional innovations and on local implementation strategies designed to facilitate access to drinking water, adapted to technical and socio-economic capacities, and to the specific cultural, environmental, political, and institutional setting. Furthermore, in the context of participation, a system should be developed where monitoring and evaluation of the solutions is carried out by the local community and not so much by the development or funding organization. Gradually the belief emerges that the vast scale of the arsenic contamination may eventually drive rural households to shift from individual to community based systems for accessing drinking water, relying on ground or surface water. This means building local drinking water institutions based on an understanding of the limitations and possible socio-economic resources. In this, the participation of the local community is believed to play an indispensable role. We believe, considering the weak central infrastructure in the area of drinking water supply, that a more efficient way to achieve proper participation is to link up with local Bangladeshi NGOs and local governmental bodies. The knowledge on community participation during a process of technological/institutional change or innovation is unfortunately very limited at that level. The existing guidelines are almost solely restricted to operation and maintenance. This paper therefore suggests a research focus on participation in the various facets of these changes. Partnerships need to be built between the local communities, development workers and scientists. Without ignoring everybody’s objectives, it is perhaps more the scientists, policy makers, development workers and funding organizations that participate in the arsenic problem, which takes place at village level. Finally, considering the urgency of the problem there should also be an initial focus on shortterm mitigation. Long-term solutions should be based on a better (scientific) understanding of the 308

causes and transportation mechanism of the contamination and the social and institutional implications. This paper aims to bring forward ideas on how to glue together the short- and the long-term objectives but first the complexity of problem will be presented.

2

BANGLADESH AND THE ARSENIC PROBLEM

The different perspectives on the arsenic problem presented in this paragraph have the objective to broadly cover the immense scope of main and related sub-problems. While it still merely scratches the surface, such a representation has the function to show how complicated the issue really is. 2.1

General background

Bangladesh is in many ways what we perceive as a typical developing country; it is predominantly agrarian, highly in debt, and with a huge disparity between rich and poor. Additionally, Bangladesh faces problems of over-population and is located in one of the wettest regions of the world. Despite that, Bangladesh achieved success in providing access to safe drinking water to more than 97% of the population, mainly through shallow hand tube wells. Now, the detection of arsenic in shallow groundwater has emerged as a new disaster. Two-thirds of the tube wells installed over the last thirty years – roughly 10 million in total – turn out to contain arsenic concentrations above the permissible levels set by the World Health Organization (Smith et al. 2000). These wells were installed in firm conviction that they would contribute to a secure and reliable drinking water supply, and would provide an alternative to the irregular surface water sources with its associated bacteriologic diseases. In itself that goal has been reached. It is therefore a bitter observation that it is this very approach that has led to widespread arsenic poisoning. The arsenic is of geological origin, but verifiable causes of the high concentration and release into the groundwater are yet to be fully understood. Arsenic is chronically toxic after prolonged low level exposure and can lead to skin lesions, bronchitis, diabetes and eventually tumors and cancers. Estimations tell that roughly 40 million people (more than 30% of the population) have been exposed to an arsenic concentration above the WHO standard for many years. Suffering most from the problem are the rural poor. They form by far the largest section of the population and at the same time have least access to societal resources. This inhibits them to take initiatives on their own. But at the same time the Bangladeshi government, seriously hampered by a lack of proper institutional and financial means can barely fulfill its normal private tasks. Added to the generally urban-biased policies of the central government it becomes clear that the rural communities have to rely on their own resources to address the problem they face. 2.2

Defining the problem field

In this light, the arsenic (chemical) contamination of groundwater as such is not the real problem; it merely aggravates the situation by rendering an already scarce and difficultly accessible resource into a health hazard. The same can be said for water-borne (biological) contamination of surface water. Obviously, unsafe drinking water resources and means to access the supplies cause the health hazards. As such this should again not be a problem within an appropriate organizational structure with the adequate institutions to tackle it. In rural Bangladesh these are however lacking. Furthermore, the main problem lies not in the mitigation technology itself, as several mitigation technologies for the arsenic problem already exist. It is their application in the current situation that is problematic and the reason for certain solutions not to sustain. It is recognized nowadays that the solutions cannot be found in a single discipline. We argue that the various technologies are appropriate for specific regions, hydrogeological and socio-economic situations, and are directly related to financial resource and institutional competence. 309

3 3.1

THREE PERSPECTIVES ON THE PROBLEM Central elements

Numerous problematic elements can be listed such as: lack of awareness, different types of contamination from different types of water sources, general water scarcity, doubts about the causes of the contamination and other research uncertainties, etc. The real issues however are found in the causal relationships between the elements and in their dynamics. Unsafe drinking water supplies, health hazards and lack of drinking water institutions can be seen as three central elements of the problem. Starting from these, three different orientations on the problem will now be exposed and the visible trends and changes discussed. In short, the first perspective, starting with the unsafe drinking water supplies, reveals a discrepancy between what is available and what is needed. It mainly relates the tangible or physical matters of the problem. Secondly, the health hazards perspective shows the relation between research experience and uncertainties and their effects on local coping mechanisms. It focuses on the knowledge system. Thirdly, the lack of drinking water supply institutions perspective relates the institutional bottlenecks with the effects on local arrangements. It is rooted at the level of organization (Rammelt 2002). Simply put, we see the problem as a situation (unsafe drinking water supply), which causes a problem (health hazard) that needs to be solved (drinking water institutions).

3.2

Physical oriented perspective

Obviously current risks are at hand for the simple reason that water is a fundamental human need and the proper means to acquire clean water are extremely limited. When the problem is considered in this way (as resulting from basic indispensable human handling), the hazards can be seen as resulting from a general scarcity of clean water, but more importantly an unequal and inefficient distribution of it. The latter is the result of rural areas being outreached and lacking in infrastructure and local technical facilities to efficiently tackle the scarcity. Another cause might be the top-down push of western inappropriate technologies. In a context such as Bangladesh, what we would define as an effective supply might be over designed and therefore inefficient, or a poor design, which might be neither effective nor efficient, but may nevertheless make a significant if sub-optimal contribution (Cairncross et al. 1980). Grossly this interpretation shows a discrepancy between what is available on a technological level and what is needed to appropriately access water resources in the specific physical conditions of rural environments. It is associated with the physical, tangible matters of the problem. Drinking water can be obtained from different sources, but each have characteristics related to quality, quantity, reliability/sustainability, cultural acceptability, costs, etc. There are broadly three types of water resources: rain, surface, and ground water. At first glance surface water is the most obvious resource due to its abundance. Unfortunately this abundance is highly irregular in many areas, and the numerous ponds and rivers are a source of water-borne diseases such as cholera, dysentery and diarrhea epidemics. Rainwater is an alternative but only during monsoon so it would require harvesting space and storage for use during the full dry season, which would be too expensive. For each source of water a number of different methods for supply are available but each have obvious disadvantages. At the same time there is an impatient drive to find the perfect solution. When we consider also non-technical factors the perfect solution is simply not available yet. This partly helps explaining why so little has been done regarding implementation. At first glance the groundwater aquifers could provide year-round safe supply. There are two types of aquifers: deep and shallow, generally separated by an impermeable clay layer. The cheapest way to access it is by using shallow ring wells that provide relatively clean water that might however get contaminated with flood water during monsoon or dry up in mid dry season. The next option is to dig deeper to the ‘young’ shallow aquifer, which turns out to be contaminated with arsenic, or the ‘older’ deep aquifer that in first indications seems to be arsenic free. Most of the 310

tested deep tube wells are essentially arsenic free but only initial observations can be made regarding the long-term. The deep arsenic free aquifers are likely to remain so, at least under natural water flow conditions (BGS/DPHE 2001). The screening of contaminated wells is the first logical step towards a solution. There are broadly two ways to do this: through lab or field-testing. It is important that the (dis-) advantages of any analysis are considered in the perspective of the purpose of the test. Studies have indicated the possibilities of seasonality in arsenic concentrations. This means that measurement at field level is not a one-off programme. Testing programmes have helped in identifying arsenic contaminated ‘hot-spots’ (Jakariya et al. 2003). However, building a supply around the safe ‘green’ wells is not possible in villages where only contaminated ‘red’ wells have been identified. (In this quick and dirty approach, wells are painted according to the contamination.) Examples from our own fieldwork experience show that wells are painted and repainted by different organizations without mutual knowledge and where a change in concentration has been measured over a period of time with the effect that people start losing their trust. It is not hard to understand that the speed of changes is mainly dictated by socio-economic factors and not by technological elements. Especially in this case where technologies are in fact available and their development path is relatively short. Moreover, technologies have to be chosen according to the exact geological and hydrological status of the area and the form and concentration of the arsenic present. While most processes show high effectiveness in laboratory studies there is little experience in actual application. Also, depending on the technology and uses, issues such as sludge disposal (a by-product of arsenic filtering processes) or excess water and drainage may cause additional risk for environmental contamination. In the short-term, a few technologies are likely to gain momentum but only because they fit certain social preconditions. Several parties show a preference for the deep tube wells, a familiar technology with familiar use. The main immediate problem with these is the higher cost and the community perceives it as a gift and not as something they should set-up and maintain themselves. As mentioned, in specific geographic units the degree of arsenic contamination and costs of the desired solutions may drive rural households to shift from individual to shared systems for accessing drinking water (van Geen et al. 2003). Household- or community-based options all have specific disadvantages but in general implementation of household-based options will be very difficult when involving thousands of families and at the same time ensuring proper use. On a household level, villagers would like to see more sustainable solutions than those the removal technologies proposed until now, which in general seem to be running only for a few months (Einwachter et al. 2001). Furthermore it will be difficult to improve/alter the technologies de-centrally when better alternatives become available. Accordingly, a community-oriented approach is not only preferred for a supply based on groundwater but also for those based on surface water, as surface water bodies are usually owned by more than one family (although surface water also needs treatment for other parameters). On the other hand, community-based solutions need to include centralized operation and maintenance by trained caretakers, which until now has rarely sustained. 3.3

Knowledge oriented perspective

General health problems and the more specific issues of coping with water contamination are largely due to poor education and/or lack of awareness. They emerge from a huge shortage of practical hands-on experience backed up by important scientific uncertainties (regarding health, geology, technology, etc) and often ambiguous, one-sided information dissemination. The interpretation shows the negative effect of unfinished debates and research uncertainties on the awareness of the community and on the urgent need to cope with the problem. This new situation was triggered by the fact that the community needed to shift from surface to ground water, which unfortunately also meant shifting from a bacteriological to a chemical contamination. In the field of research and applicability in the context of Bangladesh a general remark would be that scientific uncertainties have a huge impact on the lack of concurrent awareness as well as short-term alternatives. Regarding geology and groundwater, a study mentions that the data at 311

hand do not extend over more than a couple of years and can therefore not be taken as proof of systematic changes (BGS/DPHE 2001). The contamination is probably only apparent now because groundwater has been extensively used for drinking water in the rural areas only in the last 20–30 years. The arsenic has probably been present in the groundwater for thousands of years and in comparison, the time span of current research seems trivial. It is likely that any changes for better or for worse will be rather slow and probably take centuries. Although the overall cause of the contamination is most probably of geological origin, verifiable causes of the high concentrations are yet to be fully established. Scientists still argue on the mechanism of release. Sediments from the Himalayas carry arsenic since the start of the formation of the delta. They have gone from an oxidizing environment to a strongly reducing environment. The shallow aquifer, especially under the dry season water table (where most shallow tube wells are found) provides this condition (oxyhydroxide reduction theory) (Nickson et al. 1998, Bhattacharya et al. 2002, Ahmed 2004). While most researchers have disproved the idea that variations of the water table due to irrigation/flooding would result in oxidization of the arsenic-bearing pyrite (Pyrite Oxidation theory), there could still be a relation between irrigation and arsenic release but in another way (Harvey et al. 2002). When pumps withdraw huge amounts of water from the aquifer during the dry season, organic carbon rich waters and other chemicals are drawn down. In the aquifer this results in reducing and dissolving the iron oxides, and releasing the arsenic. In relation to agriculture a few uncertainties remain as well, such as the effect of phosphate fertilizers and other chemicals on the contamination, or the impact of flooding and the cultivation of IRRI or Boro rice (rice widely cultivated during the dry season) by minimizing air entry to the underlying aquifer and therefore intensifying any ongoing reduction and arsenic release. This process probably cannot account for the large-scale problem but nevertheless needs further investigation. The potential uptake of arsenic into root plants and food through irrigated water is also still uncertain, as well as retention in soils and leaching back to shallower aquifers, and the possible formation of arsenic-laden dust particles. Regarding health issues, there could be a release of arsenic into the water in different forms, depending on the local conditions and the applicable theory. Arsenic occurs in several oxidation stages in water, in oxidizing conditions it is dominated by arsenate (As(V)), in anaerobic conditions by arsenite (As(III)). Both arsenate and arsenite are dangerous and carcinogenic, especially the latter. This would result in different impacts on health. Health problems are linked to medical science uncertainties. Clinical symptoms of poisoning seem to vary between affected regions. One reason might be that there is no clear agreement on the definition and symptoms of arsenic poisoning. Other reasons are more directly related to the human condition. Knowledge in the relation between general health and nutritional status, and the long-term effects of arsenic and dose is incomplete. Moreover, even if external symptoms do not show, there is still no proof that there are no internal problems. Also, controversies still exists on applicability of acute poisoning medication against chronic poisoning, on reversibility of the poisoning, etc. Besides uncertainties there are also other fundamental problems with the health education system. The programmes are often unrealistic and do not account for economic restrictions in a village. Also current approaches are based on universal hygiene messages assuming that knowledge of educators is superior to the local insights and practices that have been based on years of trial and error. Finally, controversies and misunderstandings on the local level divert the discussions away from the real issues. The belief that arsenicosis might be infectious, contagious, or hereditary might be unfounded but the associated social consequences are real. In our fieldwork experiences we have seen how affected people are sometimes rejected within families, communities, or schools because the disease triggers a social stigma. The lack of awareness on the risks is the first obstacle at the local level. The confusion is immense and emerging from medical, geological, and technological (research) uncertainties. However, once the risk becomes evident as it already is in some areas; it is still not sufficient to actually trigger a similar reorganization as the one we saw thirty years back. 312

First of all, the women, generally responsible for managing family water and hygiene, have practically developed systems to supply the household of ‘clean’ water. Regarding arsenic there simply hasn’t been a chance yet to develop similar systems. Secondly, economic restrictions lie behind this phenomenon. The poor are at greater risk because they are physically more vulnerable to arsenic but also because they have fewer options for water supply and they are not in a position to ‘invest’ in prevention schemes, postponement and hoping for a cure is the only alternative. Moreover, a widespread shift would nowadays be more problematic because women’s activities around the house are already extremely time-consuming without adding the need to fetch drinking water from an alternative source. 3.4 Institution oriented perspective As mentioned earlier, the health hazard cannot be tackled without some form of a communitybased drinking water institution. History has shown that a top-down governmental approach is simply not suitable to reach to the level of a village. This has everything to do with financial shortcomings (increasing towards the bottom) and the lack of proper (multi-sectoral) policies. The difficulties for a bottom-up approach lie in the weak power of the local organizations (NGO and GO) and the problematic collaboration between them. Grossly this line shows institutional difficulties on different levels and the related bottlenecks for an organizational set-up on a local level. First of all, uncertainties have discouraged many organizations (local to international) to undertake specific actions for the affected areas. Even in our western society it is very difficult to get the authorities to react when natural disasters occur. In Bangladesh it is virtually impossible, for reasons of lack of knowledge but also of capital shortage and institutional weakness. The British colonization has had a huge and lasting impact on the system. The feeling to challenge or question it in light of new emerging situations seems to be lacking. This can hardly be blamed on a government with such a short history of independence. Gradual but noteworthy changes are however visible. Over the years a slow shift in decentralization of the government administrative units has been observed. At the bottom of the ladder, on Union Parishad (UP) level, a legacy of the British rule, some hesitating changes are visible. On one hand democratization and recently women empowerment but on the other hand also a growing corruption. In relation to the drinking water sector, the UP falls under the responsibilities of the Ministry of Local Government for Rural Development & Co-operatives (LGRD). All governmental bodies addressing rural development fall under this ministry. It is divided into different ‘thematic’ departments. The Department of Public Health and Engineering (DPHE) is responsible for supplying safe water to the community. Its organizational structure is limited down to the Upazila (sub-districts) level so that it has no direct organizational instruments to reach to a village. Most initiatives coming from that level have to be processed through the entire line up of ministers, secretaries, and officers. Eventually the decision is altered in such ways that the needy are usually not the beneficiaries. It is not surprising to see village drinking water supplies being placed in the advantage of certain influential villagers at the expense of the poorest. In specific relation to arsenic, the Bangladesh Arsenic Mitigation Water Supply Project (BAMWSP), conceptualized jointly by the Government of Bangladesh (GOB) and the World Bank – Swiss Agency for Development & Co-operation (WB-SDC), was made effective in 1999. It has accomplished a screening phase in a number of Upazilas and is working towards mitigation activities. A specific objective is to set up decentralized committees to oversee and promote project activities. Unfortunately, the community’s share in decision-making and investment is not always guaranteed. Under the Ministry of Health and Family Welfare (MHFW), the Directorate of Health on the other hand does reach the village level but does not yet have the means to tackle multidisciplinary issues and there is a lack of understanding of the links between health, food, and water. Therefore, the identification of patients does not yet trigger further investigation, such as identifying location of the village, quality of the pump, information about the family situation, nutrition, health status, 313

water usage, etc. A similar observation can be made regarding the drinking water sector, which is not included in a more general water management sector. Thus, approaches for arsenic contamination remediation are limited because they often do not account for water availability in the broader context. NGO’s on the other hand have a direct link to the people but here the problems start with GO/NGO collaboration and their differences in approach and objectives, and limited power and resources. The experience regarding the arsenic contamination and possible alternatives in the NGO sector is low. The way people are experiencing the arsenic crisis has been complicated mostly by the lack of a centralized and unambiguous awareness programme; yet another interaction (or lack of) between the community and the local and national bodies. Obviously existing institutions move only slowly. A typical illustration: after the arsenic came to the agenda, around 1998, shallow-wells were still promoted and installed in the villages for three to four years after. However, bureaucracy and the centralized character of the institutions are not the only reasons for things to move slowly. Policies and guidelines concerning water resources have to be based on research. Results from initial geological research had to be processed and double checked before anything could be presented. Also the debate on the exact threshold below which contamination is still safe is as yet undecided. Arsenic contamination in ground water and thus drinking water has been linked to various health problems but appropriate methods for diagnosing arsenic affected people and treating them are as yet poorly documented. In particular the early detection of the symptoms is critical, accentuated by the fact that there are various parallel but independent actions. In many areas the density of rural settlements, growth of rural income, and the relatively improved level of service over the last few decades have created a better environment and affordability of technologies. Presently three out of four of the wells are privately owned (Smith et al. 2000) and there is a generally growing position of the private sector also in rural areas. However this involvement is limited to current available technologies in the field of drinking water and there is not yet a pro-active attitude regarding the arsenic issue. There is a need for a more serious involvement of private sectors. 3.5

Concluding remarks

There is a direct relation between speed of implementation of certain technologies and associated risks. In the seventies the decision and popular support for shifting to tube-well technologies for extracting ground water was fast and widespread. The risk of drinking bacteriological contaminated surface water was obvious for the local, national and international community. The general situation at that time has apparently fuelled this change. Now the need for a new drinking water approach is not felt as urgent as it was in the seventies. There are several reasons for that. Especially at the local level a lack of awareness on the risks of the arsenic contamination is a main obstacle. The confusion is immense and emerges from medical, geological and technological (research) uncertainties. However, once the risk becomes evident as it already is in some areas; it is still not sufficient to actually trigger a similar reorganization as the one seen thirty years back. Institutional, cultural and economic restrictions interfere with this reform. The problem as such is not purely of a technological nature, as many technical solutions have already been developed. The difficulty is that most of these have been developed for other applications and for entirely different contexts.

4 4.1

SUGGESTED APPROACH Research orientations

Over the last ten years the faculty of Technology, Policy and Management of the Delft University of Technology (DUT) has developed contacts with NGO’s working in the rural areas (Boes 1996). They 314

led to a co-operation that nowadays is focusing on the arsenic calamity. The faculty of Geology of the University of Dhaka, Bangladesh, the faculty of Civil Engineering and Geosciences (DUT), as well as the section Soil Chemistry of the Wageningen University of Agriculture are also participating. Considering current uncertainties regarding the issue, the mid- and long-term strategy will be to minimize the risks of existing and future remedies, while incorporating the essential institutional requirements for implementation. This strategy requires research for a better understanding of the hydro-geological situation as well as the institutional possibilities and restrictions. Within the consortium ideas have been developed to relate three ‘disciplinary’ research orientations: geological, institutional and implementation research. The first direction is set towards a more scientific understanding of the behavior of arsenic and groundwater. The second looks into the preconditions for building local institutional capacity. It investigates formal and informal institutions and networks at community level. The implementation research will have strong links with both these orientations. It will explore possible frameworks for implementation of sustainable drinking water (sub-) systems, facilitating a basic development strategy for people’s participation and risks minimizing for the poor. It will aim to translate social and institutional findings into a basis to make technological choices and reversely provide usable technical recommendations for implementation of the solutions. 4.2

Short-term and long-term conflict

Such a scientific approach does not solve the immediate and urgent problem in Bangladesh. A short-term strategy has to be prescribed by activities directed to a quick drive-back of arsenic concentrations in drinking water to an acceptable level. The available technical options that can be used as a base for a safe drinking water supply are easily identified. The shallow tube well, deep tube well, pond sand filter and dug well are the most prominent candidates. As seen in earlier paragraphs, the appropriateness of these short-term solutions depends on local conditions as well as the season. On the longer term the tube well could perhaps be the better option, however a lot of research has to be done before it can seriously be implemented safely again. More importantly on the long-term is the development of institutional and organizational arrangements for local drinking water supplies that meets the needs of the most vulnerable groups in the community. This is probably a very time consuming process and socioeconomic as well as cultural factors might hamper this process severely. A conflict now arises between the urgency to solve the immediate problem and the time needed to fulfill social conditions and to resolve scientific uncertainties. One of the main dilemmas of the current situation is what strategy should have the priority. Must the emphasis be laid upon the understanding of the mechanism of the arsenic contamination so that it might be easier to come up with the best solution possible or should priority be given to mitigate the arsenic problem in order to reduce exposures of arsenic as much as possible? The question is actually how to glue together the short-term urgency and long-term sustainability. To streamline the process of the long-term research and the process of institution building it is necessary that the latter should start as soon as possible, so the time lost on later implementation of safe water options can be minimized. Scientific work and field implementation can actually strengthen each other and can be linked by using so-called ‘triggers’. These triggers will cause the community to spark off concrete discussions and actions in the villages. For example a water-testing programme in a village can be used to get the attention for the arsenic issues. In the same way an innovation, the construction of a social map, or other research input can be used as triggers. To bring about the involvement of the community will require more than just an awareness programme; it is necessary to give them a perspective on a short-term solution for their immediate problem (no survey without service). For example the installation of safe water options (such as pond sand filters, deep tube wells or rainwater harvesting systems) in the target villages might serve this purpose. This will create room for further research on long-term solutions and the knowledge and information from the local people is an essential ‘trigger’ for this. 315

5

CONCLUSIONS

In the arsenic issue two technological orientations can broadly be distinguished: the physical/ geo-chemical research as well as the technical solutions for remediation. We would like to suggest the importance of a study on the geochemistry influencing the transport mechanism of arsenic in groundwater linked to a study on available technologies and their optimization in a risk assessment. Alongside, implementation research should be started including participation issues. This does not imply that the design of purely technique-oriented solutions (such as arsenic removal filters for example) should not receive attention as well. As the available technologies have been developed for other contexts, they have to be adapted to fit the local circumstances. Nevertheless, the great majority of today’s initiatives are already focusing on technological remediation, and many technical solutions are at hand and are even being installed. The implementation of these options however happens on a limited scale and viewed from a development perspective, the long-term effects are far from clear. It is therefore our basic perception that the development of technical solutions, while very important, should not divert the attention away from implementation-oriented research. The crux of this assertion lies in the view that implementation is the main bottleneck and the reason behind the limited successes until now. It is clear that it is not the lack of technical solutions that causes the delay in solving the arsenic problem. Technical development of solutions is valuable for approaches with a short-term character, but for a sustainable development, scientific uncertainties have to be clarified. Firstly the geological/health knowledge is not ready to use and must be further developed. Secondly the implementation of the solutions will be severely hampered by the lack of good and efficient institutional support. For a successful implementation of a technology one has to realize that certain organizational conditions must first be fulfilled. Until now only minor attention has been given to the development of the necessary institutional structures. Hopefully, in the time span of a few years’ better technologies will be available. By that time it will be easier to shift to other solutions provided the institutional process would have been built up. Since the underlying processes in setting up these structures tend to have large time constants it will be almost inevitable to avoid large societal discontent during the implementation phase. Therefore a short-term mitigation programme has to be initiated as well. Ultimately there is a necessary comparison to be made in a risk assessment approach with other solutions as well, such as those based on surface water for example. Such an approach should consider the (potential) risks of existing and future technologies while also taking into account the various contaminants in groundwater or other water resources, as well as other intake sources (arsenic enters the food chain via irrigation). Only then can a field level mitigation technology be properly optimized and can a local accessible data-gathering procedure be set up to monitor the quality of the drinking water. Such field activities would in return help reduce the knowledge gaps in the current understanding of the contamination mechanism. Detailed scientific investigations could be carried-out at small-scale research sites under specific geological, hydro-geological environments with contrasting land use patterns and socio-economic conditions. Summarizing our comments we would suggest action-oriented research complementary to other, mostly technical proposals. The implementation of a ‘new’ technology is a difficult job that must be enhanced by examining the problem as a whole and by tackling it on different levels and disciplinary fronts, and by taking into account different time constants. Finally, as presented earlier, we believe technological solutions with their short-term potential can act as triggers and are therefore valuable for projects with a longer-term focus with close involvement of the local communities who eventually should be given the autonomy to shape their own development.

REFERENCES Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A., Imam, M.B., Khan, A.A. & Sracek, O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh: An

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overview. In: P Bhattacharya, AH Welch, KM Ahmed, G Jacks & R Naidu (eds.) Special Issue: Arsenic in Groundwater of Sedimentary Aquifers. Appl. Geochem, 19(2): 181–200. BGS & DPHE 2001. Arsenic Contamination of Groundwater in Bangladesh. Vol 2 Final Report, BGS Technical Report WC/00/19. Bhattacharya, P., Frisbie, S.H., Smith, E., Naidu, R., Jacks, G. & Sarkar, B. 2002. Arsenic in the Environment: A Global Perspective. In: B Sarkar (Ed.) Handbook of Heavy Metals in the Environment. Marcell Dekker Inc., New York, 147–215. Boes, J. 1996. Initial proposal, Integrated Water Management in Tangail Area, a study into the technical possibilities for local initiatives. Delft: Faculty of Technology, Policy and Management. Cairncross et al. 1980. Evaluation for village water supply planning. New York, John Wiley & Sons. Chambers, R. 1992. Rapid but relaxed and participatory rural appraisal: towards applications in health and nutrition. In: Gleason GR & Scrimshaw NS, eds. Rapid assessment procedures: Qualitative methodologies for planning and evaluation of health related programmes. Boston, MA: International Nutrition Foundation for Developing Countries. Einwachter, M., Gorp, van M. & Hilders, M. 2001. Arsenic Mitigation Strategies in North Jalirpar. Delft. Faculty of Technology, Policy and Management, Arsenic Mitigation and Research Foundation. Harvey, C.F., Swartz, C.H., Badruzzaman, A.B.M., Keon-Blute, N., Yu, W., Ali, M.A., Jay, J., Beckie, R., Nieden, V., Brabander, D., Oates, P.M., Ashfaque, K.N., Islam, S., Hemond, H.F. & Ahmed, M.F. 2002. Arsenic mobility and groundwater extraction in Bangladesh. Science 298: 1602–1606. Jakariya, M., Chowdhury, A.M.R., Hossain, Z., Rahman, M., Sarker, Q., Khan, R.I. & Rahman, M. 2003. Sustainable community-based safe water options to mitigate the Bangladesh arsenic catastrophe – An experience from two upazilas. Current Science 85(2): 141–146. Korten, D.C. 1996. Beyond Bureaucracy: The Development Agenda, Making it to the 21st Century depends on the participation of everyone and the re-creation of our major institutions, in Making It Happen, Effective Strategies for Changing the World. Adapted from IN CONTEXT #28 Kranzberg, M. 1986. Technology and History: Kranzberg’s Laws. Technology and Culture 27: 544–560. Nickson, R., McArthur, J., Burgess, W., Ahmed, K.M., Ravenscroft, P. & Rahman, M. 1998. Arsenic poisoning of groundwater in Bangladesh. Nature 395: 338. Rammelt, C.F. & Boes, J. 2002. Arsenic Mitigation and Social Mobilization in Bangladesh, Special Issue for the Engineering Education in Sustainable Development conference, Delft October 2002. Delft: Faculty of Technology, Policy and Management. Rammelt, C.F. 2002. Strategies for implementation of drinking water supplies in Bangladesh, Development by Design (dyd02) conference proceedings, Bangalore, December 2002, Bangalore. van Geen, A, Ahmed, K.M., Seddique, A.A. & Shamsudduha, M. 2003. Community wells to mitigate arsenic crisis in Bangladesh. Bulletin of the World Health Organization 81(9): 632–638. Smith, A.H., Lingas, E.O., Rahman, M. 2000. Contamination of drinking-water by arsenic in Bangladesh: a public health emergency. Bull. World Health Organization 78(9): 1093–1103.

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Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Sustainable safe water options in Bangladesh: experiences from the Arsenic Project at Matlab (AsMat) Md. Jakariya, Mizanur Rahman, A.M.R. Chowdhury Research and Evaluation Division, BRAC, Dhaka, Bangladesh

Mahfuzar Rahman, Md. Yunus, Abbas Bhiuya, M.A. Wahed Center for Health and Population Research, International Centre for Diarrhoeal Disease Research, Bangladesh (ICDDR,B), Dhaka, Bangladesh

Prosun Bhattacharya, Gunnar Jacks Groundwater Arsenic Research Group, Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Stockholm, Sweden

Marie Vahter Division of Metals and Health, Karolinska Institutet, Stockholm, Sweden

Lars-Åke Persson Department of Women’s and Children’s Health, Uppsala University, Uppsala, Sweden

ABSTRACT: The presence of elevated levels of naturally occurring arsenic in groundwater of Bangladesh, has severely impaired the decade long effort of providing safe water to nearly 98% of its population and putting an estimated 35 million people-nearly one fourth of the total population – at risk. In order to address this problem, a project titled “Arsenic in tubewell (TW) water and health consequences in Matlab Upazila of Chandpur district (AsMat)” is being implemented jointly by ICDDR,B and BRAC. During this study. all the TWs in Matlab have been assigned unique identification numbers, with marked GPS coordinates, depth, and age. It is estimated that nearly 65% of the about 13,000 TWs in Matlab have As concentrations above the Bangladesh drinking water standard (50 ␮g/L). In order to minimize arsenic exposure, a work to provide various alternate safe drinking water options to the exposed population has been initiated. As of March 2004, about 1047 different alternate safe water options, such as Pond Sand Filter (PSF), Rainwater Harvester (RWH) and different filters to remove arsenic as well as pathogenic bacteria, were distributed among the targeted exposed population in Matlab. To ensure sustainable use, the provided options were assessed based on community acceptability, technical viability, and financial viability.

1

INTRODUCTION

The arsenic problem in Bangladesh has been described as the largest mass poisoning in history (Rahman et al. 2001, Smith et al. 2000) Chronic ingestion of inorganic arsenic causes characteristic skin lesions as observed in populations in Bangladesh and elsewhere. The presence of arsenic above the Bangladeshi limit of safe drinking water (50 ␮g/L, WHO 2001) was first detected in groundwater of the Bengal Delta Plain (BDP) aquifers in Bangladesh in 1993 (Bhattacharya et al. 1997, 2002, 2004, Nickson et al. 2000, Smedley & Kinniburgh 2002, Ahmed et al. 2004). This has resulted in a severe environmental disaster affecting several million people in the region, as groundwater is the main source of potable water for nearly 98% of the population in Bangladesh. What was a success story is now poised to threaten the lives of millions of people living in 61 out of the 64 districts in Bangladesh (Fig. 1). Given the present scenario of safe water status, it would 319

Figure 1. Table 1.

Distribution of arsenic in Bangladesh (DPHE/BGS/DFID 2000). Five major challenges for arsenic mitigation in Bangladesh.

• Test all tube wells for arsenic; there are about 8–10 million such wells in Bangladesh; • Investigate the mechanisms of arsenic contamination of groundwater; • Identify people with elevated risk of arsenic toxicity. A variety of toxic effects, including different types of cancers (most of which are not likely to be reversible); • Provide alternative arsenic-safe water options which are culturally acceptable, technically feasible, environmentally safe, and affordable by common people; • Research into the extent of soils contamination due to irrigation with arsenic-contaminated groundwater that may affect the food chain.

not be possible for Bangladesh as well as other arsenic affected parts of the globe to ensure the target of the Millennium Development Goals of ensuring safe drinking water supply to its inhabitants by the year 2015, if necessary measures are not taken on an urgent basis to alleviate the crisis. Though elevated As levels in the groundwater in Bangladesh were detected in the early 1990’s, it has only received adequate attention since 1998. The Government of Bangladesh, non-governmental organizations (NGOs) and donors are working together to address this critical issue. The As contamination in the groundwater of Bangladesh poses five major challenges for action (Table 1). The most daunting challenges among the above are those of testing of the wells, finding alternative sources of safe water, and identification of people at highest risk of arsenic exposure. BRAC in conjunction with ICDDR,B and collaborating institutions in Sweden, is implementing a project at Matlab upazila of Bangladesh to address arsenic and its remedial issues. This project is benefiting from ICDDR,B’s Health and Demographic Surveillance System in 142 villages in Matlab. The surveillance system contains demographic information, reproductive outcomes, health and nutritional information, which are linked with a geographic information system. The project activities started in 2001 and will continue till 2006. Some preliminary findings on the mitigation are presented here. 320

Figure 2.

2 2.1

Location map of the AsMat study area.

MATERIALS AND METHODS The AsMat study area

Matlab is situated 53 km south east of Dhaka, accessible by road and river transport (Fig. 2). The location of Matlab is highly affected by the sedimentation process of As laden soil, as it is situated near the Meghna River, where it joins the confluent streams of the Brahmaputra and Ganges rivers. The area is a low-lying delta plain intersected by branches of the rivers and numerous canals. During the monsoon most of the land is flooded, except for clusters of houses built on earthen mounds. In 1988–89, a 60 km long embankment was built alongside the bank of the Dhonagoda and Meghna rivers. The embankment was built primarily to protect the area from monsoon flooding so that agricultural activities might be carried out throughout the year. ICDDR,B is running a health and demographic surveillance system in 142 villages of the Matlab upazila, encompassing a 220,000 population in 18,386 hectares of land. 2.2

Screening of TWs

Geographical coordinates of all the identified TWs in AsMat study area were recorded using a handheld Global Positioning system (GPS) unit. Testing TW water to identify TWs, with low arsenic concentrations (less than the Bangladesh drinking water standard, 50 ␮g As/L), is the first step in reducing the magnitude of As contamination for the millions of exposed population in the country. To date, more than a million wells in Bangladesh have been tested using field kits (Rahman et al. 2002). The wells with As levels ⬎50 ppb were painted red to indicate that the water is unsafe for drinking, and those with levels ⬍50 ppb were painted green to indicate that the water is safe. However, it should be pointed out that arsenic concentrations 50 ppb still possess a very high cancer risk, compared to what is recommended by the WHO (WHO, 2001) for an environmental pollutants such as arsenic. In the Matlab study area, a total of 13,302 TWs were tested using the 321

Merck field kit and were numbered and classified as to geographical coordinates by Geographic Positioning System (GPS). Field teams consisting of a field research assistant from ICDDR,B and Sasthya Sabika (Health workers) from BRAC performed the testing of TW water for As (Chowdhury & Jakariya 1999). Earlier, they were given a one-week extensive training on how to test TW water and on different deliverable messages regarding arsenic to be disseminated among villagers after testing of TW water. Tubewells were purged adequately (20/30 strokes) prior to testing with the Merck field kit and collecting a sample for laboratory analysis. One Merck field kit is capable of performing ca. 80 tests and costs around $50, giving an approximate price of $0.63 per test. The Merck kit uses a semi-quantitative colorimetric method. Both arsenate and arsenite species present in the water samples are chemically converted to arsine gas, which reacts with a mercury bromide impregnated test paper stripe and produces a color that correspond to the amount of As present in the sample. 2.3

Community involvement

Although the presence of elevated levels of As in groundwater has been known since the beginning of 1993, there is a substantial need to raise the awareness among the population about the health risks associated with long-term ingestion of contaminated groundwater through drinking. Awareness raising and motivation on As issues were done in two ways. First, people are made to understand that As is highly toxic, and second, people are told that they should procure safe water at their own effort and cost. A combination of approaches such as staging popular theatre, meetings, organizing workshops, and showing video are being used to raise awareness about As contamination and the consequences of drinking As contaminated water (Hossain et al. 2003). Sustainable development cannot be achieved without involving the local community in the planning and development processes. The local inhabitants were involved in implementing the project. Multiple meetings at the village level were held at different stages of the project to inform and involve the community. At these meetings potential safe water options were discussed. The villagers decided where the community-based systems would be best located and then committed to maintaining the system. 2.4

Providing alternative safe water options

Installation and assessment of safe water options were major activities of the mitigation component of the project. Obviously, arsenic free water is the main way to reduce the risk of arsenic toxicity (Jakariya et al. 2003). It may be noted that the options installed in Matlab are based on BRAC’s previous experience of providing alternative safe water options to other arsenic exposed populations of the country. Specific criteria were used while targeting households for alternative options (Table 2). The next priority, in terms of providing safe water options, has been given to the villages where 50% or more of the TWs were found to be contaminated with As. Within the village, families with poor socioeconomic condition were given priority for distribution of home-based options. The community-based options are constructed in places where most of the TWs are contaminated. It was decided that the alternative options distributed in the project areas should not be free of cost. The community or the individual (in the case of household-based options) should have to contribute at least 20% of the total installation cost, either in cash or in kind, according to the inclination of the individuals. It was observed that the community preferred to pay their contribution part Table 2. • • • •

Criteria for distribution of the safe water options.

Willingness to bear 20% of the installation cost and 100% operation and maintenance cost Water As concentrations in the village Symptoms of arsenic toxicity in the village Poor socioeconomic condition

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in kind (e.g. paying the carrying cost of construction materials, arranging lodging facility for the mesons during construction period, etc), particularly for the construction of PSF and RWH system. 2.5

Monitoring of water quality

The water quality of the provided options was monitored both for arsenic and bacterial contamination. A monthly plan has been developed to check water quality of the randomly selected options before and after treatment of water to see the technical viability of the provide options. 3 3.1

RESULTS AND DISCUSSION Arsenic mapping

Out of the total 15,884 TWs in the project area, only 13,302 functioning TWs were identified and tested for As contamination using field kits. About 65% of the total tested TW waters indicated As concentrations at levels exceeding the Bangladeshi permissible limit (i.e. 50 ppb) and about 9% of the total TWs tested had As levels over 500 ppb. With the help of these GIS maps, highly As contaminated pockets were efficiently targeted for distribution of alternative safe water options. This piece of information was useful to target not only the vulnerable areas but also to demonstrate safe water options to the wider community for their own understanding as well as for decision making with respect to the best suitable options for themselves (Fig. 3). 3.2

Arsenic concentration in relation to depth and age of TW

An attempt was made to assess whether there was any relationship between depths of TWs and the level of As contaminations as it was well known that As concentration decreases with depth (JICA 2002, BGS 2000). As shown in Figure 4, there was a statistically significant association between depth of TWs and the As concentration (p ⬍ 0.01). More than 14% of the TWs within 17 m depths were found to have low arsenic concentrations. As contamination was found to be highest (more than 95% TWs) within the depth range of 17–67 m, and almost 100% TWs were found to be within

Figure 3. Distribution of TWs and their status of As concentration in groundwater in a village of AsMat study area.

323

% 100 90 80 70 60 50 40 30 20 10 0

5680

249

611 67 248 0-16

17-33

66

3

33-50 Depth (m)

As-safe (0- 50 µg/L)

Figure 4.

186

899

405

51-67

> 67

As-contaminated (>50 µg/L)

Variation of As concentration with depth of the tubewells.

Bangladeshi drinking water standard (BDS) beyond 67 m deep (Fig. 4). The conclusion that can be drawn from these results is that the probability of getting water with low As concentration is at depths more than 67 m in this area. Large-scale installation of TWs in Bangladesh was initiated in the early 1980s, when people decided to install their own TW for access to safe and clean water or to gain prestige. However, almost 82% of the present TWs in Matlab were installed during the last 10 years (Fig. 5). As shown in Figure 5, As concentrations are higher in older TWs. 3.3

Alternative water option

The main options promoted (BRAC 2000, Jakariya et al. 2003) were: (i) treatment of surface water with community-based Pond Sand Filter (PSF) and home-based Bishuddhya filter; (ii) collection of rainwater in Rain Water Harvesters (RWH); (iii) treatment of As contaminated groundwater with home-based three-pitcher filters, Alcan filters and Safi filters (Table 3) The number of total exposed families in Matlab is about 26,650. About 54% of the exposed families obtained water from the green marked TW and another 13% were drinking water from different types of alternative water options already installed/distributed in Matlab. The remaining families were to take advantage of evaluations of existing options in their areas. Before testing TW water for As contamination in Matlab, it was observed that about three families shared one TW. The number of families covered by green marked TWs were calculated based on this statistic. But recently we conducted a survey in several villages of Matlab with the aim to know the number of users of green marked TWs (As-free). According to the survey, the number of users of green TW has increased from three to seven, which indicates most of the exposed families are at least trying to collect water from different available safe water sources, primarily from the green marked TWs. The options presented in Table 3 are being assessed on several criteria: initial and running costs; ease of implementation, running and maintenance; continuity and flow of supply; As removal capacity; susceptibility to bacteriological contamination; and acceptability to the community. It is not yet possible to draw general conclusions as to the sustainability of the provided options. A description of the different options provided in Matlab is presented in the following. 3.3.1

Description of the options

3.3.1.1 Pond sand filter (PSF) In the coastal belt of Bangladesh, where salinity in the shallow groundwater is reported to be a severe problem, the Department of Public Health Engineering (DPHE) and UNICEF have installed 324

%

100 90 80 70 60 50 40 30 20 10 0

1914 3129

771 133 0-10

11--20 Age of TWs (Years)

No. of As-safe TW (0-50 µg/L)

Figure 5.

43

170 21+

No. of As-contaminated TW (>50 µg/L)

Percentage of arsenic contaminated TWs with the year of installation.

Table 3.

Safe water options distributed in Matlab Upazila during the AsMat Project.

Safe water options Household based options Three pitcher filter Alcan filter Safi filter Bishudhha filter Rainwater harvester (RWH) Community-based options Pond sand filter (PSF) Total

Total number of options 99 653 24 140 115 16 1047

Total families covered 99 653 24 140 115 1,600 (100/PSF) 2631

slow sand filtration units linked. Hand pumps are used to deliver pond water to the units, which are called Pond Sand Filters (PSF). In this slow sand filtration system a bed of fine sand is used through which the water slowly percolates downward, resulting in the removal of pathogens through a combination of physical and biological processes. One PSF can supply the daily requirement of drinking and cooking water for about 40–60 families with a cost of about US$ 600. Once trained, masons with locally available materials can construct PSF. There is no chemical treatment involved in this process and little effect on the environment. The greatest challenge for this option is to find suitable ponds which are perennial, free from pisciculture, and protected from using for bathing and washing clothes, cattle, etc. It has been reported that many of the PSFs constructed in the past in other arsenic affected areas where BRAC is providing As mitigation options, are now abundant as the owners’ of the ponds restarted commercial fishing activities (BRAC 2000). Most ponds are privately owned and therefore, villagers have little to say about changing owners’ decision about restarting commercial fishing activities in spite of having formal agreement with the community that they will reserve the ponds for drinking water use only. In Matlab, it can be noted that because of this type of practical problems only 16 PSFs have been constructed during the last two years. However, there is potential that PSFs could be one sustainable alternative to the arsenic problem in some areas. 325

3.3.1.2 Rain water harvester (RWH) Rainwater harvesting is utilised in many parts of the world to meet the demand for drinking water. There is a long-established tradition of rainwater collection in some parts of Bangladesh, where shallow groundwater water has elevated salinity. Although the potential for rainwater harvesting is good in some areas of Bangladesh, the amount of rainfall varies widely across the country. Rashid (1977) shows that mean annual precipitation ranges from 1,400 mm (about 55 inches) along the country’s central western border to more than 5,000 mm (200 inches) in the far northeast. The wet months are mid-June to late September and the dry period is from January to April. About 80% of the annual precipitation occurs in the monsoon period. Rainfall patterns were also confirmed with local communities in order to ascertain the feasibility of RWH, and alternatives and parallel use of other options were considered before constructing RWH jars in Matlab. The capacity of a jar is about 32,000 litres and the cost is about US$ 130 (DPHE/UNICEF 1988–93). During the course of the project it was observed that the construction cost was too prohibitive to construct for an individual rural household. Monthly analysis for total coliform and faecal coliform bacteria to assess the safety level of the RWH water for drinking was done at laboratory and the results were found to be highly satisfactory. 3.3.1.3 Safi filter This household level filtration device was developed locally by Professor Sayeed Safiullah of Jahangirnagar University, Dhaka, Bangladesh. The Safi filter was designed to remove both As and pathogenic bacteria. The filters distributed initially had many problems in terms of flow rate and As removal efficiency (BRAC 2000). The Safi filter consists of two concrete buckets of 20 litre storage capacity, one of which is placed inside the other. The upper bucket is filled with TW water, which then flows through a permeable ‘candle’ and is collected in the lower bucket where it is stored. When needed it is drawn off with a tap. One small Safi filter is designed to filter approximately 40 litres of water per day. The cost of such filter is US$ 15. The Safi filter candle is prepared from a chemical mixture of laterite soil, ferric oxide, manganese di-oxide, aluminium hydroxide and meso-porous silica. These materials adsorb arsenic as the water passes through the candle and thus the contamination is removed. It is also claimed that the candle eliminates pathogenic bacteria from the contaminated water. The manufacturers calculated that after two years of continuous use the candle would need to be replaced with a fresh one. Each new candle costs US$ 4. 3.3.1.4 Bishuddhya filter This non-chemical based filter is basically designed to remove pathogen load from surface water. The filtration and purification technique used in this system is similar to PSF technology. Bishuddhya filters are relatively inexpensive and produced in Bangladesh. The main material used in this filter is different mesh sizes of locally available rocks and the water passes from the bottom through different layers to remove bacteria before it arrives at a storage chamber. The cost for this householdbased device is US$ 45 with virtually no operation and maintenance cost, except for washing the materials after certain interval depending on the suspended loads of row water. The plastic container of this filter is made from food-graded plastic. The only precautionary measure that needs to be ensured here is that surface water must be obtained from a protected source to ensure no contamination with chemicals, fertilisers, etc. This filter could be an ideal option for areas where protected surface water is available and for southwestern part of the country where most people still use surface water for drinking and cooking purposes. 3.3.1.5 Three-kolshi or three-pitcher filter The three-pitcher filter is based on an indigenous method of filtration, which has been used in Bangladesh for ages. Traditionally, two local clay pitchers (called kolshi) are used to filter water. The top pitcher is partially filled with sand and charcoal, and a small hole is made in the bottom. A piece of synthetic cloth is placed over the hole to prevent sand from spilling out. Water is passed through this pitcher to remove suspended matter from surface water and more recently to remove iron from TW water. After passing through the top pitcher, filtered water is stored in the bottom 326

pitcher. Scientists from Bangladesh and the USA noted the potential of this simple method to remove As well. They modified the system by adding a third pitcher above the sand/charcoal pitcher, which is filled with iron filings to provide an additional source of iron oxide to adsorb more As (Rasul et al. 1999). The total cost for developing such a unit is less than US$ 5. There are, however, a number of unanswered questions about the three pitcher filters that need to be addressed before they are taken up on a larger scale. First, bacteriological contamination of the water occurs before or during filtration. It has not been proven beyond doubt that the water from the three-pitcher filter will be completely free of chemical impurities. The potential for trace elements such as lead, chromium, zinc, tin, etc. to enter the water from the iron filings must be conclusively discounted. The final question about the three-pitcher filter is whether it is technically effective in the long run since it has been observed that after three/four months of continuous use the filters start leaching As. However, the three-pitcher system has enormous potential to provide an emergency drinking water source for the As-affected areas in rural Bangladesh, if the problems mentioned above are taken care of. 3.3.1.6 Activated alumina filter (ALCAN filter) The basic principle of this system is adsorption of As by activated alumina. In this filter the raw water passes upward through the activated alumina media and the treated water becomes As-free. Activated alumina is formed by the thermal dehydration (250 to 1150°C) of an aluminium hydroxide such as, gibbsite, bayerite, etc. Its principle characteristic is high surface area (⬎200 m2/g) and associated porosity. The term activated refers to the capacity of the alumina to enter into adsorption and/or catalytic reactions, and is determined largely by such variables as crystal structure, pore size and distribution, and the chemical nature of the surface. There are two types of ALCAN systems available: household based units and community-based units. It has been observed in Matlab that household based units are more preferred by people than that of the community-based units. The activated alumina used as media in the system has to be imported. The initial cost is high for both type of units: US$ 260 per unit (unit ⫹ media), and US$ 52 (unit ⫹ media) for the household based unit. Running costs are required because the activated alumina needs to be changed periodically. The replacement cost of media for community-based unit is US$140 to treat 80,000 litres of water and for household based unit is US$ 12 to treat 11,000 litres of water. Apart from the cost which seems prohibitive for both individual and community members to replace media on their own, disposal of used material is also another issue of concern not only for Alcan filter but also for all other arsenic removal filters available in Bangladesh. These two issues need to be monitored properly to avoid further hazard.

3.4 Community views about the options In order to assess peoples’ opinions about the provided options, several focus group discussion sessions (FGD) were carried out in different parts of Matlab. The following matrix is prepared based on the information collected using the FGD technique. Table 4 shows the ratings of each of these factors on a scale of 1 to 5. The maximum possible score is 45 and a higher rating indicates more potential. From the above exercise, it can be assumed that community preference for PSF was highest among all other available technologies in spite of the few PSFs constructed in Matlab. This small number may indicate the unavailability of ideal ponds required for such construction. On the other hand, it is also true that villagers were not habituated to use PSFs in these areas like people of the coastal belt areas where surface and rainwater harvesters are the only sources for drinking water. It can also be concluded from Table 4 that the rating of PSF, RWH, ALCAN, and Bishuddhya filters were fairly close. In Matlab, when the provided options were assessed to identify a pattern of distribution, interestingly it was observed that the options followed a unique criterion: PSF, where individual or community was found to be ready to sacrifice their earnings from pisciculture and also from other domestic uses; Bishuddhya filters, in areas where most of the villagers were found 327

Table 4.

Preference of the alternative safe water options.

Parameters Initial cost Running costs Ease of implementation Technical effectiveness Maintenance required? Monitoring required? Continuity of supply Susceptibility to bacteriological contamination contamination Social acceptability Total

PSF

RWH

Alcan filter

3-pitcher filter

Bishuddha filter

Safi filter

1 4 3 4 4 4 4

2 5 3 3 4 3 2

2 1 4 5 3 2 5

5 3 2 4 2 1 4

3 4 4 4 3 3 4

2 3 2 1 1 1 2

3 5 32

4 4 30

4 4 30

2 2 25

2 3 30

3 1 16

to be still using surface water for drinking and cooking purposes after alum treatment or simply boiling; on the other hand, relative well-off villagers and also the people who do not want to move from using TW water were found to be adopted Alcan filters. 4

CONCLUSIONS AND RECOMMENDATIONS

Supply of As free, safe drinking water remains the most crucial issue in the As mitigation programme. The geohydrological situation of Bangladesh varies from area to area. Similarly, soil types and water chemistry vary from region to region. Water chemistry is particularly important in developing As removal technologies because performance of a number of As removal technologies depends on water quality, especially pH. For immediate safe drinking water supply, short-term options such as As removal technologies can be used. However, to solve the problem permanently, a long-term solution needs to be adopted. In an area where the proportion of TWs having elevated As concentrations is not high, people can often collect water from the nearby As-safe TWs. Sludge disposal generated from the As removal technologies are an issue of worries. If not disposed of properly, it may have the potential to pollute the surrounding safe ground water again or create a new environmental hazard. It involved much effort and time to convince people to shift from using relatively polluted surface water to safe ground water. Now people have a strong preference for TW and it would not be easy to shift these large populations from the TW option again. The situation might be worse in areas where no symptoms of arsenic toxicity have become visible even after a long duration of As exposure. Therefore, a strong Behavioural Change Communication (BCC) component should be added with the water supply programmes in order to bring positive changes in the behaviour of the users population. Before promoting any safe drinking water options, it should be tested at the community level with an adequate sample size because many water options including As removal technologies found to be effective under ideal conditions or at the laboratory may not function properly at the community level. Any water supply options should be assessed in terms of acceptability, affordability and accessibility for the users. The community should bear the responsibility for proper operation and maintenance, without which the water options may rapidly fail. For each water option, caretakers need to be identified from among the users and trained on operation and maintenance issues. Hygiene education is also very important as water may get contaminated not only at the source but also at the distribution, collection and storage point. Another important issue that has not gained adequate importance so far is the health aspect of any given water options. In order to mitigate the As problem, is the water option increasing the risk of any other water borne or water related diseases? Thus risks and benefits analysis of any As mitigation options should be assessed. 328

Finally, it is clear that the technologies introduced in this project to supply As free safe drinking water are only short-term emergency solutions for areas severely affected by As contamination. The longer-term solutions may include the provision of piped water supply to its population and the optimum use of its surface water.

ACKNOWLEDGEMENTS The authors would like to thank Swedish International Development Cooperation Agency (Sida) for supporting the project titled “Arsenic in tubewell (TW) water and health consequences in Matlab Upazila of Chandpur district” undertaken by ICDDR,B, and BRAC. We are thankful to Richard Johnston, UNICEF, Bangladesh for providing his valuable suggestions on an earlier draft of this manuscript.

REFERENCES Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A., Imam, M.B., Khan, A.A., & Sracek, O. 2004. Arsenic enrichment in groundwater of the alluvial aquifers in Bangladesh: An overview. Appl. Geochem. 19(2): 181–200. Bhattacharya, P., Chatterjee, D. & Jacks, G. 1997. Occurrence of arsenic contaminated groundwater in alluvial aquifers from Delta Plains, Eastern India: Options for safe drinking water supply. Int. Jour. Water Res. Management 13: 79–92. Bhattacharya, P., Frisbie, S.H., Smith, E., Naidu, R., Jacks, G. & Sarkar, B. 2002. Arsenic in the Environment: A Global Perspective. In: B. Sarkar (ed) Handbook of Heavy Metals in the Environment Marcell Dekker Inc., New York, pp. 147–215. Bhattacharya, P., Welch, A.H., Ahmed, K.M., Jacks, G. & Naidu, R. 2004. Arsenic in Groundwater of Sedimentary Aquifers. Appl. Geochem. 19(2): 163–167. BRAC 2000. Combating a deadly menace. Early Experiences with A Community-based Arsenic Mitigation Project in Rural Bangladesh. Res. Monogr. Ser., 16: 1–116. Chowdhury, A.M.R. & Jakariya, M. 1999. Testing of Water for Arsenic in Bangladesh. Science 284(5420): 1621. Hopenhayn-Rich, C., Browning, S.R., Hertz-Picciotto, I., Ferreccio, C., Peralta, C. & Gibb, H. 2000. Chronic arsenic exposure and risk of infant mortality in two areas of Chile. Environmental Health Perspectives 108(7): 667–673. DPHE and UNICEF 1988–93. “Water Supply and Sanitation Project Implementation Guideline and Specification for the Rural Areas of the Coastal Region”. DPHE/BGS/DFID 2000. Groundwater Studies of Arsenic Contamination in Bangladesh, Final Report, Dhaka. Hossain, Z., Quaiyum, M. & Jakariya, M. 2003. Using materials for mass communication: experiences of an arsenic mitigation project in Bangladesh. Bangladesh Journal of Mass Communication and Publishing, 2(1). Jakariya, M., Chowdhury, A.M.R., Hossain, M.Z. & Rahman, M. 2003. Management of chronic arsenicosis patients in Bangladesh. Tropical Doctor 33(4): 251–252. Jakariya, M., Chowdhury, A.M.R., Hossain, Z., Rahman, M., Sarker, Q., Khan, R.I. & Rahman, M. 2003. Sustainable community-based safe water options to mitigate the Bangladesh arsenic catastrophe – An experience from two upazilas. Current Science 85(2): 141–146. JICA/DPHE 2002. “The groundwater development of deeper aquifers safe drinking water supply to arsenic affected areas in western Bangladesh”, Final report, Dhaka. Khan, A.W. and Ahmad, SK.A. 1997. Arsenic in drinking water: Health effects and management. A training manual. Department of Occupational and Public Health. National Institute of Preventive and Social Medicine (NIPSOM), Dhaka. Nickson, R.T., McArthur, J.M., Ravenscroft, P., Burgess, W.G. & Ahmed, K.M. 2000. Mechanism of arsenic release to groundwater, Bangladesh and West Bengal. Appl. Geochem. 15(4): 403–413. Rahman, M.M., Chowdhury, U.K., Mukherjee, S.C., Mondal, B.K., Paul, K., Lodh, D., Biswas, B.K., Chanda, C.R., Basu, G.K., Saha, K.C., Roy, S., Das, R., Palit, S.K., Quamruzzaman, Q. & Chakraborti, D. 2001.

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Chronic arsenic toxicity in Bangladesh and West Bengal, India—a review and commentary. J. Toxicol. Clin. Toxicol. 39: 683–700. Rahman, M.M., Mukherjee, D., Sengupta, M.K., Chowdhury, U.K., Lodh, D., Chanda, C.R., Roy, S., Selim, M., Quamruzzaman, Q., Milton, A.H., Shahidullah, S.M., Rahman, M.T. & Chakraborti, D. 2002. Effectiveness and reliability of arsenic field testing kits: Are the million dollar screening projects effective or not? Env. Sci. Technol. 36: 5385–5394. Rashid, H. 1977. The Geography of Bangladesh. University Press Limited, Dhaka. Smedley, P.L. & Kinniburgh, D.G. 2002. A review of the source, behavior and distribution of arsenic in natural waters. Appl. Geochem. 17: 517–568. Smith, A.H., Lingas, E.O. & Rahman, M. 2000. Contamination of drinking-water by arsenic in Bangladesh: a public health emergency. Bull. World Health Organization 78 (9): 1093–1103. WHO 2001. Arsenic in drinking water: Fact Sheet 210. Geneva. http://www.who.int/mediacentre/ factsheets/fs210/en/print.html (Accessed on March 9, 2004).

330

Natural Arsenic in Groundwater: Occurrence, Remediation and Management – Bundschuh, Bhattacharya and Chandrasekharam (eds) © 2005, Taylor & Francis Group, London, ISBN 04 1536 700 X

Prerequisite studies for numerical flow modeling to locate safe drinking water wells in the zone of arsenic polluted groundwater in the Yamuna sub-basin, West Bengal, India Soumyajit Mukherjee Department of Earth Sciences, Indian Institute of Technology Roorkee, Roorkee, Uttaranchal, India

ABSTRACT: Arsenic beyond permissible limit from tube-well water has been reported from Yamuna sub-basin, which was initially giving potable water. Locating safe water tube-wells using MODFLOW software can give long-term remedy. Hydrostratigraphy of the study domain is conceptualized up to a depth of 140 m and three principal sediment groups were defined viz., Clay, Sand, and Sand and Gravel. This study indicates that the arsenic concentration in sediment and groundwater does not reveal depth-wise trend. Deviation Factor (DF) indicates irregular trend of concentration levels in sediment as well groundwater depth-wise. Statements on acquisition of arsenic free groundwater are site-specific. These necessitate modeling based decision making for safe well location. Secondly, all the DF values are negative, which means that arsenic in sediments is depleted with respect to their world average values. To locate safe drinking water wells in the Yamuna sub-basin, the modeler should incorporate depth specified hydro-stratigraphy and field constraints. This will be followed by calibration of other hydrological parameters to simulate observed arsenic spreading responsible for its sporadic nature.

1

INTRODUCTION

In the Bengal Basin, in has been reported that even the deeper aquifers have been contaminated with arsenic above the permissible limit (Muralidharan 1998, Mukherjee et al. 1999; Burgess et al. 2000, Bhattacharya et al. 2002a,b, Ahmed et al. 2004). The importance of numerical modeling in locating deep aquifers free from arsenic contamination and optimum pumping rates to extract water from such aquifers have been stressed by several workers (Mallick & Rajagopal 1996; Burgess et al. 2000, Majumdar et al. 2002). The groundwater flow model using MODFLOW (McDonald & Harbaugh 1988) can be used for this purpose (Majumdar et al. 2002). It is advantageous over other softwares because it has easy to defend codes, easy to update features, and has the facility of adding external modules (Kresic 1997, Herzog et al. 2003). The prerequisites to develop such a groundwater flow model are (i) 3D visualization of hydro-stratigraphy (Mallick & Rajagopal 1996, Vries 1997, Zhang & Brusseau 1998, Weight & Sonderegger 2001, Majumdar et al. 2002, Herzog et al. 2003) in terms of uniform layers of sediment sheets (Majumdar et al., 2002); (ii) depth wise arsenic concentration (Burgess et al. 2000) and its diffusion coefficient/leachability factor; (iii) set of boundary conditions and aquifer parameters and (iv) calibrated hydraulic conductivity of individual layers in three directions, leakance of aquitards, and contaminant dispersivity in three-directions and its molecular diffusion rate (cf. Kresic 1997, Majumdar et al. 2002). Decision variables in this exercise would be well locations, optimum pumping rates, and water withdrawal with arsenic ⬍0.01 ppm, which can be worked out with global optimization technique, or using particle tracking method with MT3D, both of which can be interfaced with MODFLOW. 331

Figure 1. Location Map showing the study area (modified after Majumdar et al. 2002). Blocks under study domain: 1. Chakdah, 2. Haringhata, 3. Bangaon, 4. Barackpore, 5. Amdanga, 6. Habra-I, 7. Habra-II, 8. Gaighata, 9. Baduria, 10. Swarupnagar. Blocks 1 & 2 belong to Nadia district, rest of them belong to North 24 Parganas district.

The present work aims to: (i) establish MODFLOW compliant hydro-stratigraphy, and (ii) to finding out depth wise trends in the variation of arsenic concentration in sediments in the Yamuna Sub-basin (a part of the Bengal basin, India). 2

THE STUDY AREA

The geographic boundaries of the Yamuna sub-basin (latitude: 22°49⬘–23°03⬘N; longitude: 88°24⬘–88–51⬘E) are demarcated by the south flowing Bhagirathi river on the west and Ichamati river on the east (Fig. 1). This sub-basin covers an area of 1500 km2 and lies both in Nadia and North 24 Paraganas districts of West Bengal and falls under the domain of Gangetic delta. The Gangetic delta comprises numerous partial fining upward sequences, characteristic of laterally shifting meandering river course (Bhattacharya et al. 1997). 3

HYDRO-STRATIGRAPHIC CHARACTERIZATION

For modeling purpose, it is difficult to simplify the hydro-stratigraphy as “aquifers separated by aquitards” in a deltaic system such as the Yamuna Sub-basin (Freeze & Cherry 1979, Premchitt & 332

Table 1. A typical borehole lithology from the Yamuna Sub-basin (for referencing, lithounits have been numbered from bottom to top) (Source: CGWB). Unit

Lithology

Depth (m)

Thickness (m)

24 23 22 21 20 19 18 17 16 15 14 13 12 11 10 09 08 07 06 05 04 03 02 01

Light brown silty clay Grey fine sand Grey medium & fine sand Grey fine & uniform sand Dark grey fine sand & mica Dark grey fine sand Grey fine sand Grey fine sand & sandstone Grey fine sand Grey fine sand & mica Grey medium uniform sand Dark grey silty clay & sand Grey fine sand Grey coarse and medium sand Grey very fine dirty sand Grey coarse sand Grey fine to coarse sand Grey fine sand Grey fine dirty sand & mica Grey sand & sandstone Grey fine sand Dark grey silty clay Dark grey silty sand Grey coarse sand & gravel

0–3.96 3.96–16.2 16.2–19.2 19.2–22.3 22.3–28.3 28.3–31.4 31.4–37.5 37.5–40.5 40.5–43.6 43.6–46.6 46.6–49.7 49.7–52.7 52.7–56.4 56.4–68.0 68.0–75.9 75.9–80.2 80.2–83.2 83.2–86.3 86.3–89.3 89.3–93.9 93.9–106.4 106.4–109.1 109.1–110.6 110.6–116.7

3.96 12.19 3.05 3.05 6.09 3.05 6.10 3.04 3.05 3.05 3.05 3.05 3.65 11.59 7.92 4.27 3.05 3.04 3.05 4.57 12.5 2.74 1.53 6.09

Das Gupta 1981). Well-to-well correlation in such a situation is difficult due to frequent vertical repetition and lack of horizontal continuity of litho-units (Fitts 2002). Hydro-stratigraphic conceptualization of the geologic domain, however, remains a prerequisite for developing any groundwater flow model (Zhang & Brusseau 1998). Extensive lateral migration of meandering rivers during the Quaternary Period within the study area has created vertical repetition and horizontally discontinuity of sediments (clay, sand and gravel). In such a situation, for modeling purpose (using MODFLOW) the hydro-stratigraphy can be simplified by (i) neglecting the narrow lenses of lithounits (Marsily et al. 1978), or (ii) grouping together the litho-units with similar hydraulic conductivity (Martin & Frind 1988, Herzog et al. 2003). Such simplification may lead to zigzag hydro-stratigraphic pattern between the wells. Subsequent to the development of litho-log based model, top and bottom elevations within the MODFLOW grid can be matched and adjusted to simulate stratigraphic conditions (Jones et al. 2002). In this work, 34 borehole lithological sections obtained from the Central Ground Water Board of India, and those obtained from literature (Bhattacharya et al. 1997; Mukherjee et al. 1999) were utilized. All these boreholes, drilled to depths varying from 25 to 400 m, partially penetrate the upper sedimentary formations. A large part of the subsurface formations are unconsolidated sediments and only at places, due to overburden, diagenesis of sand into sandstone is recorded thus resulting in reduction in permeability. For the sake of modeling, sedimentary sequences (Table 1) from the study area are grouped into three broad litho-units viz. (i) clay (including soft, hard, compact, sandy and silty clays), (ii) sand of all size ranges, and (iii) sand and gravel (Table 2). The lower limit for these groups is fixed at 140 m due to paucity of large data beyond this depth from the boreholes. In general, borehole lithological correlation is carried out using certain marker beds, which are characterized by physical, chemical and mineralogical parameters. In the present case, however,it is not possible to select any marker bed due to repetitive occurrence of similar lithological units (Table 1). 333

Table 2.

The reclassified lithological groups.

New unit

Old units

Litho-group

Recalculated depth (m)

Recalculated thickness (m)

6 5 4 3 2 1

24 14 to 23 13 4 to 12 2 to 3 1

Clay group Sand group Clay group Sand group Clay group Sand and gravel group

0–3.96 3.96–49.68 49.68–52.73 52.73–106.37 106.37–110.64 110.64–116.73

3.96 45.72 3.05 53.64 4.27 6.09

Table 3. Conceptualized multi-aquifer-aquitard system in the study domain upto 140 m depth. Hydro-stratigraphic units

Depth (m)

Thickness (m)

Clay group aquitard Sand group aquifer Clay group aquitard Sand group aquifer Clay group aquitard Sand & gravel group aquifer

0–25 25–42 42–70 70–90 90–100 100–140

25 17 28 20 10 40

A three-tier multi-aquifer-aquitard system was conceptualized and is shown in Table 3. The depths assigned to the hydro-stratigraphic units are the most frequently encountered depths in the reclassified lithological groups. It must be mentioned here that lithostratigraphic sequence based on a single borehole (see Bhattacharya et al. 1997) is site-specific and thus cannot be used for such modeling purpose.

4

VARIATION IN ARSENIC CONCENTRATION WITH DEPTH

Depth wise variation of arsenic concentration in groundwater (Table 4) and sediments (described later in ‘deviation factor’) of the study domain (and some also from other parts of the Bengal basin) were compiled and compared by the author (Mukherjee 2002). No agreement in terms of the depth range of maximum arsenic concentration in sediments and groundwater was found in these data. Moreover, generalized statements by various workers (see Table 4) were found true only for their own study domains, hence they must not act as guidelines to locate safe water tubewells within the Yamuna sub-basin, nor in any other part of the Bengal basin. Similarly, arsenic content in sediments do not show any systematic relationship with depth. Further, in order to compare enrichment or depletion of arsenic in a repetitive sedimentary sequence ‘Deviation Factor’ (DF), a unitless number, is defined as: (1) where Ca ⫽ concentration of arsenic in sediment; Cw ⫽ global average concentration of arsenic in that sediment. A positive DF indicates arsenic enrichment while a negative DF indicates arsenic depletion. A zero value indicates the concentration to be equal to the world average. Since all the lithounits in the Yamuna sub-basin are fresh water sediments, concentration of arsenic in fresh water sediments from the literature (Ghosh & Chakravorty 1996, Welch et al. 1998, Anawar et al. 2002) are compiled in Table 5, and are used to calculate DF. Average arsenic 334

Table 4.

Arsenic concentration in groundwater in the study area and other parts of the Bengal Basin.

Salient observations

Reference

Nadia district: 73.5–90 m depth: Groundwater As 200 ␮g/L Nadia district: 24–120 m depth: high arsenic zone (⭐550 ␮g/L) Ghetugachi (Nadia): 290 ␮g/L arsenic at 116 m depth Bengal Basin: Groundwater below 60 m practically arsenic free Nadia: Temporal increase of arsenic conc. in tube-well water Bengal Basin: ‘wells more than 10 m below the As enriched aquifers can pump groundwater relatively free of arsenic for some years’ Bengal Basin: Groundwater ⬍100–150 m depth are As contaminated N 24 Pargana: groundwater between 70–150 m depth are As contaminated Nadia: arsenic contaminated water from 14–109 m depth Ghetugachi (Nadia): below 115 m, arsenic conc. in groundwater 1.0 to 0.16 ppm Nadia: “arsenic concentration in sediments and groundwater showing antithetic relationship” Baruipur area: Sand beds with clay caps with more than 30 m thickness contains safe water. Yamuna sub-basin: arsenic conc. in groundwater first increases with pumping, then falls Nadia: high arsenic concentration in groundwater restricted within 35–46 m

Ghosh & Chakravorty (1996) Bhattacharya et al. (1997) Jain (1997) Saha et al. (1997) Muralidharan (1998) Burgess et al. (2000) Smedley & Kinniburgh (2002) Basu & Sil (2003) Bandyopadhyay (2002) Bandyopadhyay (2002) Bandyopadhyay (2002) Pal et al. (2002) Majumdar et al. (2002) Smedley (2003)

Note: Published data do not show correlation between the depth and levels of arsenic contaminated groundwater. Table 5.

Average arsenic concentration in fresh water sediments

Lithology

Average conc. of arsenic (ppm)

Sand Clay Shale Sandstone Claystone

4.80 12.00 13.00 1.50 3.00–10.00

(Source: Ghosh & Chakravorty 1996, Welch et al. 1998, Anawar et al. 2002). Table 6.

Depth wise variation of DF in samples from Itina, N 24-Pargana district.

Lithology

Depth range (m)

Calculated DF

Grey silty clay Dark grey clayey silt Dark grey micaceous fine to coarse sand Grey clay Reddish brown sand Light brown grey fine to coarse sand

1.60–1.64 4.48–4.52 6.28–6.32 13.15–13.30 34.70–34.90 36.15–36.30 40.15–40.30 48.15–48.30

NC NC ⬍⫺98.96 ⫺33.33 ⫺89.58 ⫺76.67 ⫺79.13 ⬍⫺98.96

NC: Not calculated.

concentration for mixed sediments (e.g. silty clay) is not available in the literature, hence DF values for these sediments have not been calculated. The DF values (Tables 6 and 7) calculated are negative thus indicating arsenic depletion. Also there is no apparent relationship between the depth and arsenic concentration in the sediments. 335

Table 7.

Arsenic concentration in sediments and DF from the study area.

Area

Lithology

Depth (m)

Calculated DF

Raw data source

N 24 Pargana N 24 Pargana Nadia Ghetugachi (Nadia) Habra (N 24 Pargana) N 24 Pargana Birohi (Nadia)

Clay Clay Sand Clay Sand Sand Sand

0–25 75–115 64–90 0–3 48.00–48.15 40.15–40.30 119

⫺93.17 to ⫺99.17 ⫺54.17 to ⫺95.42 ⫺95.77 ⫺94.81 ⫺88.33 ⫺82.5 ⫺75.00

Saha et al. (1997) Saha et al. (1997) Bhattacharya et al (1997) Jain (1997) CGWB CGWB CGWB

5

OTHER CONSTRAINTS

For developing a sound MODFLOW model, besides the conceptualized hydro-stratigraphy, the following information are essential: (i) physical boundaries of the study area, its hydraulic head boundaries; (ii) conductance values on the MODFLOW grids; (iii) quantitative water budget of the area in terms of rainfall, irrigation return flow, drawdown etc.; (iv) aquifer parameters (transmissivity and storativity values) obtained from pump test data. All these parameters have been compiled by Majumdar et al. (2002).

6

CONCLUSIONS

In the present study, the entire sedimentary sequence has been grouped into three broad lithological units as discuss above. The arsenic concentration in these sediments does not show any significant variation with depth and hence can be generalized for modeling purpose. The negative DF values indicate depletion of arsenic in the present sequence compared to the world average. There is no significant relationship between the arsenic concentration and depth wise distribution of various sedimentary sequence. Thus any attempt to model using MODFLOW to locate sites for arsenic free groundwater will be a futile exercise.

ACKNOWLEDGEMENTS The author thanks S.S. Srivastava (IITR), P.K. Majumdar and N.C. Ghosh (NIH, Roorkee), P.J. Martin (University of Waterloo), P.A. Macfarlane (Kansas Geological Survey) and B.L. Herzog (Illinois State Geological Survey) for their help in developing this model. CGWB, Kolkata, NIH, Roorkee, and D. Muralidharan (NGRI, Hyderabad) are thanked for providing data sets for this work. Critical comments from an anonymous reviewer considerably improved the manuscript. The author acknowledges the help received from Prof. D Chandrasekharam (IITB) and Dr. Prosun Bhattacharya (KTH) during the preparation of the manuscript.

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Author Index

Aguilera-Alvarado, A.F. 255 Acharyya, S.K. 17 Aguayo, A. 125 Ahmed, K.M. 31, 283 Alam, S. 119 Armienta, M.A. 77, 125

Hadjinicolaou, J. 237 Haque, N. 247 Hasan, M.A. 31 Hossain, M.A. 163 Hough, R.L. 109 Huq, S.M.I. 95, 119

Berner, Z.A. 3 Bethune, D. 67 Bhattacharya, P. 41, 57, 133, 319 Bhiuya, A. 319 Boes, J. 307 Brömssen, M. von 133 Bulbul, A. 119 Bundschuh, J. 57

Islam, M. Amirul 163 Islam, Md. N. 173

Cano-Aguilera, I. 247, 255 Ceniceros, N. 125 Chandrasekharam, D. 25 Cheburkin, A.K. 205 Choi, H. 103 Choudhury, M.S. 119 Chowdhury, A.M.R. 319 Claesson, M. 57 Cole, J.M. 67 Companywala, M.T. 273 Cruz, O. 125 Deb, P.C. 273 Driehaus, W. 189 Fagerberg, J. 57 Fletcher, T. 109 Gani, M.O. 163 Gbadebo, A.M. 85 Genç-Fuhrman, H. 223 González-Acevedo, Z.I. 205, 255 Gutiérrez-Valtierra, M. 247

Jacks, G. 41, 133, 319 Jakariya, Md. 319 Jigami, H. 297 Kanel, S.R. 103 Karim, M.A. 163 Kawai, S. 119 Khan, M.H. 31 Kim, K.W. 103 Kop
r iva, A. 49 Krachler, M. 205 Lee, M.K. 155 Leonardi, G.S. 109

Rahman, M. 319 Rahman, M. 319 Rammelt, C.F. 307 Rodriguez, R. 77 Rodríguez, R. 125 Routh, J. 145 Ryan, M.C. 67 Saiduzzaman, A.K.M. 237 Saraswathy, A. 145 Saunders, J.A. 155 Schuiling, O. 223 Shah, B.A. 17 Sharma, A.K. 263 Sharma, S. 273 Shotyk, W. 205 Sikder, A.M. 31 Singh, T.S 211 Smith, S. 67 Sracek, O. 49, 57, 133 Srivastava, A. 273 Storniolo, A.R. 57 Stüben, D. 3

Mango, H. 125 Martin, R.A. 57 McConchie, D. 223 Mejia, Gómez, J.A. 77 Misra, K. 273 Mohammad, S. 155 Moon, S.H. 103 Morrison, G. 247 Mosbæk, H. 163 Mukherjee, S. 331 Mulligan, C.N. 337

Tandukar, N. 41 Thir, J.M. 57 Tjell, J.C. 263

Naidu, R. 95

Yunus, Md. 319

Ongley, L.K. 125

Zeman, J. 49

Pant, K.K. 211 Perrusquía, G. 247 Persson, L.Å. 319

339

Vahter, M. 319 Valero, A.A. 41 Villaseñor, G. 125 Von Bernuth, R.D. 173 Wagner, F. 3 Wahed, M.A. 319