Long-Term Performance of Permeable Reactive Barriers, Volume 7 (Trace Metals and other Contaminants in the Environment)

Trace Metals and other Contaminants in the Environment 7 Long-term Performance of Permeable Reactive Barriers Trace ...

Author: K.E. Roehl | T. Meggyes | F.G. Simon | D.I. Stewart

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Trace Metals and other Contaminants in the Environment 7

Long-term Performance of Permeable Reactive Barriers


Series Editor:

Jerome O. Nriagu Department of Environmental and Industrial Health School of Public Health University of Michigan Ann Arbor, Michigan 48109-2029 USA

Other volumes in this series: Volume 1: Volume 2: Volume 3: Volume 4: Volume 5: Volume 6:

Heavy Metals in the Environment, edited by J.-P. Vernet Impact of Heavy Metals on the Environment, edited by J.-P. Vernet Photocatalytic Purification and Treatment of Water and Air, edited by D.F. Ollis and H. A1-Ekabi Trace Elements - Their Distribution and Effects in the Environment, edited by B. Markert and K. Friese Metals, Metalloids and Radionuclides in the Baltic Sea Ecosystem, P. Szefer Bioindicators and Biomonitors: Principles, Concepts and Applications, edited by B.A. Markert, A.M. Breure and H.G. Zechmeister

Cover illustration: Isolines of uranium concentration in p~g/1 in and around the experimental permeable reactive barrier in P~cs, Hungary (see Chapter 10)


Long-term Performance of Permeable Reactive Barriers

Edited by K.E. Roehl

Department of Applied Geology, Karlsruhe University, Germany T. Meggyes

Federal Institute for Materials Research and Testing (BAM), Berlin, Germany F.-G. Simon

Federal Institute for Materials Research and Testing (BAM), Berlin, Germany and D.I. Stewart

School of Civil Engineering, University of Leeds, UK

2005

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© 2005 Elsevier B.V. All rights reserved. This work is protected under copyright by Elsevier B.V., and the following terms and conditions apply to its use: Photocopying Single photocopies of single chapters may be made for personal use as allowed by national copyright laws. Permission of the Publisher and payment of a fee is required for all other photocopying, including multiple or systematic copying, copying for advertising or promotional purposes, resale, and all forms of document delivery. Special rates are available for educational institutions that wish to make photocopies for non-profit educational classroom use Permissions may be sought directly from Elsevier's Rights Department in Oxford, UK: phone (+ 44) 1865 843830, fax (+44) 1865 853333, e-maih [email protected]. Requests may also be completed on-line via the Elsevier homepage (http://www.elsevier.com/locate/permissions). In the USA, users may clear permissions and make payments through the Copyright Clearance Center, Inc., 222 Rosewood Drive, Danvers, MA 01923, USA; phone: (+ 1) (978) 7508400, fax: (+ 1) (978) 7504744, and in the UK through the Copyright Licensing Agency Rapid Clearance Service (CLARCS), 90 Tottenham Court Road, London W lP 0LP, UK; phone: ( + 44) 20 7631 5555; fax: ( + 44) 20 7631 5500. Other countries may have a local reprographic rights agency for payments. Derivative Works Tables of contents may be reproduced for internal circulation, but permission of the Publisher is required for external resale or distribution of such material. Permission of the Publisher is required for all other derivative works, including compilations and translations. Electronic Storage or Usage Permission of the Publisher is required to store or use electronically any material contained in this work, including any chapter or part of a chapter. Except as outlined above, no part of this work may be reproduced, stored in a retrieval system or transmitted in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, without prior written permission of the Publisher. Address permissions requests to: Elsevier's Rights Department, at the fax and e-mail addresses noted above. Notice No responsibility is assumed by the Publisher for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions or ideas contained in the material herein. Because of rapid advances in the medical sciences, in particular, independent verification of diagnoses and drug dosages should be made. First edition 2005 Library of Congress Cataloging in Publication Data A catalog record is available from the Library of Congress. British Library Cataloguing in Publication Data A catalogue record is available from the British Library. ISBN: 0-444-51536-4 O The paper used in this publication meets the requirements of ANSI/NISO Z39.48-1992 (Permanence of Paper). Printed in The Netherlands

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Contents

Preface Contributors About the editors Acknowledgements 1.

.

,

xi XV

xvii xix

Permeable Reactive Barriers A Introduction B Concept of permeable reactive barriers C Reactive materials for contaminant attenuation D Application and long-term performance of PRBs 1 Elemental iron barriers 2 Other reactive materials E Outlook References

5 9 9 16 19 19

Construction Methods of Permeable Reactive Barriers A Introduction B Construction of cut-off walls 1 Cut-off wall types 2 Cut-off wall materials C Construction of reactive barriers 1 Design considerations 2 PRB construction technologies 3 Cost analysis 4 Outlook Acknowledgements References

27 27 27 28 32 34 35 38 48 48 49 49

Materials and Processes for Uranium Removal from Contaminated Water A Introduction B Materials and experimental procedures C Attenuation processes 1 Zeolites 2 Hydroxyapatite 3 Activated carbon

53 53 53 55 55 59 67

1 1 1

vi

Contents

4 Hydrated lime 5 Elemental iron 6 Iron oxides D Summary and conclusions References Appendix 3A Behaviour of Uranium in Elemental Iron and Hydroxyapatite Reactive Barriers: Column Experiments A Introduction B Initial laboratory column test systems 1 Materials and methods 2 Results of the laboratory column experiments C Experiments with radiotracers 1 Materials and methods 2 Experimental results D Conclusions References Appendix 4A

o

,

67 68 71 72 73 76

77 77 81 81 83 86 87 90 99 100 104

Laboratory Tests Using Natural Groundwater A Introduction B Column experiments 1 Experimental set-up 2 Experiments on shredded cast iron 3 Experiments on waste steel fibres 4 Experiments on hydroxyapatite C Floor-scale tests 1 Design and operation 2 Experiments on shredded cast iron 3 Geochemical modelling of mining water in contact with elemental iron 4 Experiments on waste steel fibres 5 Experiments on hydroxyapatite 6 Long-term performance of the reactive materials D Conclusions References Appendix 5A Appendix 5B

111 111 111 111 114 114 115 116 116 119

On-site Column Experiments A Introduction B Columns in monitoring wells 1 Experimental set-up 2 Results

137 137 138 138 138

121 124 125 127 129 130 131 132

Contents

C

Large-scale field column experiments 1 Experimental set-up 2 Results D Conclusions Acknowledgement References

°

New Barrier Materials: The Use of Tailored Ligand Systems for the Removal of Metals from Groundwater A Introduction B Concept and development 1 Calixarenes 2 Salens 3 Polymers C The preparation of PANSIL D Efficiency of contaminant attenuation 1 Mechanism of contaminant removal by PANSIL 2 Efficiency and durability of PANSIL 3 Comparison tests with ion exchange resins 4 Comparison of PANSIL performance with that of the ion exchange resins 5 Other factors likely to effect the performance of PANSIL E Technological applicability F Conclusions Acknowledgements References Electrokinetic Techniques A Introduction B Scope and approach C Experimental set-ups and methods 1 Electrokinetic cell 2 Aquarium 3 Model soils and reactive material 4 Model solutions D Theoretical model E Results 1 Small-scale experiments (electrokinetic cell) 2 Bench-scale experiments F Discussion and conclusions 1 Summary of the experimental results 2 Practical aspects G Outlook References

vii 139 140 143 150 151 151

153 153 154 154 154 155 155 156 157 160 172 178 178 178 180 181 181 183 183 185 186 186 188 189 190 190 193 193 200 203 203 205 207 207

viii

Contents

Mecsek Ore, P6cs, Hungary Case Study A Historical overview B Waste characterisation 1 Characterisation of the waste rock piles 2 Wastes from heap leaching 3 Tailings from conventional milling C Monitoring 1 Tailings ponds area (Site I) 2 Heap Leaching Pile II and Waste Rock Pile I (Site II) 3 Valley below Waste Rock Pile III (Site III) 4 Mill site and Heap Leaching Pile I (Site IV) D Site characterisation, site selection 1 Geological setting of the contaminated sites 2 The geological environment of the tailings ponds 3 Geological structure of the southern area of Shaft I (Waste Rock Pile III) 4 Screening and ranking of the sites E Detailed investigation of Sites II and III 1 Geophysical investigation 2 Hydrological investigation of Site III F Conclusions Acknowledgements References Appendix 9A Appendix 9B Appendix 9C

227 227 229 229 232 242 242 242 243 253 257

10.

Experimental Iron Barrier in P6cs, Hungary A Introduction B Design of the permeable reactive barrier C Construction phase D Results of operation 1 Water chemistry 2 Hydraulic performance of the PRB E Conclusions References

261 261 261 262 267 267 272 280 281

11.

Installation and Operation of an Adsorptive Reactor and Barrier (AR&B) System in Brunn am Gebirge, Austria A Introduction B General description 1 Location and case history 2 Geological-hydrogeological set-up 3 Contamination of the unsaturated zone 4 Groundwater contamination

283 283 284 284 285 285 286

.

211 211 215 215 216 217 218 219 221 222 223 224 224 226

Contents

Site assessment 1 Environmental and hydrogeological investigations 2 Evaluation of potential environmental hazards 3 Legal aspects D Concept of project implementation 1 Site-specific preconditions 2 Evaluation of technical variants 3 Planning of project implementation 4 Groundwater protection E Assessment of system- and site-specific suitabilities 1 Feasibility study 2 Suitability of the permeable reactive barrier concept F AR&B system implementation 1 Hydraulic barrier 2 In situ adsorptive reactors and runoff system 3 Adsorbent reactive material G System operation - hydraulics and water chemistry 1 Groundwater flow induced by the AR&B system 2 Monitoring of groundwater chemistry and purification efficacy 3 Environmental effects on site conditions 4 Routine monitoring 5 Advanced monitoring 6 Permeability of the reactive material in the reactors 7 Tracer tests to describe the hydraulic function of the reactors 8 Investigations on reactive material H Perspectives and outlook References

286 286 288 289 289 289 289 290 290 291 291 292 292 292 294 294 295 295

Regulatory and Economic Aspects A Introduction B Regulatory aspects 1 Technology-inherent concerns 2 Approval of permeable reactive barriers for groundwater remediation C Economic aspects 1 PRB costs 2 Reactive materials 3 Economic aspects of the experimental barrier installation in P6cs, Hungary D Outlook References

311 311 311 311

C

12.

Index

ix

296 299 299 299 301 302 303 307 308

315 316 316 318 319 320 320 323

This Page Intentionally Left Blank

Preface

Groundwater is one of mankind's most important natural resources because it is the main source of drinking water. Contaminated sites resulting from industrial activity, mining, improper waste disposal or accidents involving hazardous substances pose a permanent threat to aquifers. Harmful chemicals can leach from polluted areas, for instance through rain water infiltrating the soil, and migrate downwards until they reach an underlying aquifer. The groundwater may become contaminated as a result and no longer usable as drinking water. It is, therefore, very important to develop and implement methods of preventing and reducing groundwater pollution. Pump-and-treat, the most frequently used conventional method for groundwater remediation, exhibits a number of shortcomings, while permeable reactive barriers (PRBs) represent a new and innovative technology with many advantageous features. PRBs enable physical, chemical or biological in situ treatment of contaminated groundwater by bringing it into contact with reactive materials. The reactive material is inserted underground in a natural aquifer and intercepts the pollution plume as it is carried along within the aquifer, and thus the contaminants are treated without either wholesale soil excavation or water pumping. This cost-effective clean-up technology has much less impact on the environment than other methods, and since it requires hardly any energy input during operation, it is generally more economical over the long term than methods such as pump-and-treat that require continuous energy input. While extensive research has been performed on many technological aspects of PRBs, and a number of contaminants have already been successfully treated by PRB systems, long-term performance has not been extensively considered and little is known about the processes influencing the long-term behaviour. This gap in our knowledge is all the more problematic because design life has a decisive influence on the economic viability of PRBs. This book describes methods for the evaluation and enhancement of the long-term performance of PRB systems, especially of those targeting heavy metals such as uranium and organic contaminants, by sorption and/or precipitation mechanisms. The contents originate mainly from original research work performed within an international collaborative project funded by the European Commission. The project was called "Long-Term Performance of Permeable Reactive Barriers Used for the Remediation of Contaminated Groundwater" (acronym: "PEREBAR") and was undertaken between the years 2000 and 2003. Processes that impair the barrier performance during PRB operation, and technologies to enhance the long-term efficacy of PRB systems were studied qualitatively and quantitatively. Two case study sites formed the central part of the project. The primary case study site was the former Hungarian uranium ore mining and processing

xii

Preface

area near the city of Prcs in Southern Hungary. The second site is located in Brunn am Gebirge, Austria, where an activated carbon PRB system is installed to treat a plume of organic contaminants at a former industrial site. The first two chapters of this book introduce the field of PRBs as an innovative technology for passive groundwater remediation. Chapter i gives a brief introduction to the concept of PRBs. Potential reactive materials and the major biogeochemical mechanisms that can be utilised in PRBs are presented, and design considerations discussed. Particular attention is given to an up-to-date review on the application of PRBs, especially on the experiences and lessons learnt about the long-term performance of fullscale installations. Chapter 2 describes the practical aspects of PRB construction. In the first part of this chapter, cut-off wall construction methods are described, since PRB construction techniques are based firmly on experience gathered with these methods and because cut-off walls are integral parts of some PRB designs (e.g. funnel-and-gate systems). In the second part, PRB design considerations and construction techniques (including some innovative techniques) are explicitly discussed. The next four chapters focus on the removal of uranium from contaminated groundwater by a selection of reactive materials including zeolites, hydroxyapatite (HAP) and elemental iron (Fe ~ also widely referred to as "zero-valent iron" (ZVI)). The driving force behind the investigations described in these chapters was the case study site in Prcs, Hungary. Therefore the experimental conditions develop, chapter by chapter, towards the actual field conditions at that site. Chapter 3 describes the results of batch experiments conducted to evaluate the effectiveness of natural zeolitic tuff, hydroxyapatite, activated carbon, hydrated lime, elemental iron and iron oxides in removing uranium from aqueous solution. The experiments were conducted with simple solutions of uranyl nitrate dissolved in deionised water. It was found that elemental iron and hydroxyapatite are the most effective materials in removing uranium from water. Therefore, further experiments focused on these two materials. Column experiments described in Chapter 4 showed that elemental iron and hydroxyapatite have strong uranium attenuation capabilities. The column experiments also showed the susceptibility of elemental iron to corrosion effects and the formation of secondary mineral precipitates due to the extreme geochemical conditions inside the iron matrix. A special feature of the column experiments was the development of a nondestructive method to measure the propagation of the uranium front during an experiment using a radiotracer. The column experiments were conducted with an artificial groundwater with a composition close to that of the case study site in Prcs. Chapter 5 describes the set-up and results of laboratory experiments conducted with groundwater taken from monitoring wells at the site in Prcs. These experiments focussed again on elemental iron and hydroxyapatite, and were conducted as column experiments and floor-scale box experiments (at cubic meter scale). The results of the experiments showed that the composition of the local groundwater, with its high concentrations of Ca, Mg, HCO3 and SO4, has a significant impact on the long-term functioning of PRB systems based on elemental iron. Laboratory-based experiments are usually subject to rather artificial boundary conditions, and a number of crucial parameters and groundwater constituents may differ significantly from real field conditions. Therefore, with the experiments described in Chapter 6, the step from the laboratory to the field was taken. On-site column experiments

Preface

xiii

with elemental iron and hydroxyapatite were designed and operated at a field test location on the former Hungarian uranium ore-processing site near the city of Prcs. The test site was located downstream of a large uranium-bearing waste rock pile where the shallow local groundwater showed significantly elevated concentrations of dissolved uranium. To simulate different flow conditions, small-scale columns were installed and operated within monitoring wells, and large-scale column experiments were located and operated very close to the monitoring wells used to supply their influent groundwater. The results of these on-site (and even partly in situ) column experiments showed that both elemental iron and hydroxyapatite effectively remove uranium from contaminated groundwater under field conditions. For elemental iron it could also be shown that the residence time of the contaminated groundwater in the reactive material controls, to some extent, the change of overall groundwater composition. It can be concluded that the volume of the reactive zone and the groundwater flow-velocity through the reactive zone are important design parameters for controlling adverse effects that occur within the elemental iron barriers (such as precipitation of secondary minerals) and are thus important to the long-term performance and operating life of iron-based PRBs. The next two chapters address innovative ideas on improving the performance of PRBs, especially with respect to their long-term behaviour. Chapter 7 describes the development and testing of a new reactive material designed to sequester uranium (VI) from contaminated groundwater. This material, named PANSIL, consists of a polymeric resin coated onto the surface of quartz sand. The advantage of PANSIL is that it is selective towards the uranyl ion, available in a granular and durable form (important considerations for materials to be used in a PRB system), and does not have any side-effects on either the groundwater composition or the geochemical conditions in the barrier, and thus avoids secondary effects such as coating or clogging of the reactive matrix. Chapter 8 evaluates the feasibility of electrokinetic methods to positively affect the long-term efficiency of PRBs. The approach of coupling electrokinetic processes with PRB systems to reduce the advective transport of groundwater constituents that may impair the barrier function was studied in a series of laboratory experiments. The results described in this chapter suggest that the installation of an electrokinetic fence upstream of the barrier could indeed electrokinetically trap such groundwater constituents outside the barrier. Chapters 9-11 report investigations and research work conducted at the two case study sites in Hungary and Austria. The characteristics of the former Hungarian uranium ore mining and processing site near the city of Prcs in Southern Hungary, including its pollution history, hydrogeology and ongoing remediation activities, are presented in Chapter 9. The chapter also describes the site investigations carried out to find and characterise a suitable location for an experimental PRB system to treat the uraniumcontaminated groundwater. In Chapter 10, the design, construction and operation of an experimental elemental iron barrier at the Prcs case study site are described. The aim of this pilot-scale barrier is the removal of dissolved uranium from the local groundwater. It consists of 38 tonnes of shredded cast iron placed in a shallow aquifer in a small valley downstream of a large, uranium-bearing waste rock pile. During the first year of operation uranium removal from the groundwater by the experimental barrier has been very successful. At the same time the change in overall groundwater composition of the barrier effluent indicated that strong geochemical processes were taking place inside the barrier material. These processes include the formation of significant amounts of precipitates,

xiv

Preface

mainly carbonates, which in the long-term might lead to changes in the hydraulic properties of the system. Chapter 11 describes the geological, hydrogeological and environmental characteristics of a former industrial site in Brunn am Gebirge, Austria which is heavily contaminated with organic contaminants such as polycyclic aromatic hydrocarbons (PAH), hydrocarbons, BTEX, phenols and chlorinated hydrocarbons. A groundwater remediation scheme, called a "Adsorptive Reactor and Barrier (AR&B) System", has been implemented at the site. This system is designed as a hydraulic barrier with four gates that funnel the groundwater through adsorptive reactors containing activated carbon. This chapter also describes the routine monitoring applied at the site to document the groundwater clean-up efficiency of the AR&B system and additional sampling and testing conducted at the activated carbon reactors to investigate those hydrogeochemical parameters that may allow an assessment of the long-term performance of the system. In the final chapter, Chapter 12, some issues concerning the regulatory acceptance of PRBs and cost data currently available for PRBs are discussed. By addressing the issue of long-term performance of PRBs, an important aspect of this technology, we aim to advance PRB technology as an accepted, scientifically sound, costeffective and stable tool for passive groundwater remediation. Thus we hope to contribute, with this book, to an improvement in pollution management and a reduction in the exposure of groundwater resources to harmful pollutants, and thereby safeguard water resources for future generations.

Karl Ernst Roehl Karlsruhe, Germany Tamrs Meggyes, Franz-Georg Simon Berlin, Germany D. L Stewart Leeds, United Kingdom

Contributors

Catherine S. Barton Mineral Industry Research Organisation, Leeds, LS1 2ES, UK (formerly at the University of Leeds)

Mecsek4rc Environmental Protection Co., Eszterg~r Lajos u. 19, H-7633 P6cs, Hungary

Zsolt Berta

Vera Biermann Federal Institute for Materials Research and Testing (BAM), Department IV, Environmental Compatibility of Materials, Unter den Eichen 87, D-12205 Berlin, Germany

Department of Process Engineering, Miskolc University, Egyetemvfiros, H-3515 Miskolc, Hungary

J6zsef B6hm

School of Chemistry, University of Leeds, Woodhouse Lane, Leeds, LS2 9JT, United Kingdom

David E. Bryant

J6zsef Csicsdk Mihdly Cs6vdri

MECSEKI~RCRt., Eszterg~ir Lajos u. 19, H-7633 P6cs, Hungary MECSEK-0KO Rt., Eszterg~ir Lajos u. 19, H-7633 P6cs, Hungary

Department of Applied Geology (AGK), Karlsruhe University, Kaiserstrasse 12, D-76128 Karlsruhe, Germany

Kurt Czurda

Department of Mining and Geotechnics, Miskolc University, Egyetemv~ros, H-3515 Miskolc, Hungary

Akos Debreczeni

Hansj6rg Fader

Fader Umweltanalytik, Reichardtstrasse 30a, D-76227 Karlsruhe,

Germany Gdbor F61ding MECSEK-0KO Rt., Eszterg~r Lajos u. 19, H-7633 P6cs, Hungary

Department of Process Engineering, Miskolc University, Egyetemv~ros, H-3515 Miskolc, Hungary

Imre Gombk6t6

Gabi Gregolec Department of Applied Geology (AGK), Karlsruhe University, Kaiserstrasse 12, D-76128 Karlsruhe, Germany


Petra Huttenloch

xvi

Contributors

Department of Mining and Metallurgical Engineering, Laboratory of Metallurgy, National Technical University of Athens, GR-157 80 Zografos, Athens, Greece

Athina Krestou

Stefan Ludwig

Fader Umweltanalytik, Reichardtstrasse 30a, D-76227 Karlsruhe,

Germany Chris McDonald School of Civil Engineering, University of Leeds, Woodhouse Lane, Leeds, LS2 9JT, United Kindgom Tamrs Meggyes Federal Institute for Materials Research and Testing (BAM), Department IV, Environmental Compatibility of Materials, Unter den Eichen 87, D- 12205 Berlin, Germany

School of Earth and Environment, University of Leeds, Woodhouse Lane, Leeds, LS2 9JT, United Kingdom Katherine Morris

Gruppe Umwelt + Technik GmbH (G.U.T.), Leonfeldnerstrasse 18, A-4040 Linz, Austria

Manfred Nahold

Ingenieurkonsulent fiir Technische Geologie, Weidlinger Stral3e 14/3, A-3400 Klosterneuburg, Austria

Peter Niederbacher


Dimitris Panias


Ioannis Paspaliaris


Karl Ernst Roehl

Federal Institute for Materials Research and Testing (BAM), Department IV, Environmental Compatibility of Materials, Unter den Eichen 87, D-12205 Berlin, Germany Franz-Georg Simon

Gdtbor Simoncsics

MECSEKI~RC Rt., Eszterg~ir Lajos u. 19, H-7633 Prcs, Hungary

School of Civil Engineering, University of Leeds, Woodhouse Lane, Leeds, LS2 9JT, United Kindgom

Douglas L Stewart


Anthimos Xenidis

About the editors

Karl Ernst Roehl is a senior lecturer at the Department of Applied Geology at Karlsruhe University where he graduated in Geology and also obtained his Ph.D. in Applied Geology (1997). Following a 2-year postdoctoral stay at the Hebrew University of Jerusalem, Israel, his current activities include teaching and research in clay science and environmental geology, and participation in national and international research and cooperation projects in these fields. Tamfis Meggyes graduated in petroleum engineering from Miskolc University, Hungary. Since then, Dr Meggyes has dealt with fluid mechanics, hydraulic transport of solids, jet devices, laser Doppler velocimetry, landfill liners, groundwater remediation and tailings facilities in Hungary, UK, Germany and USA. He has published over 100 papers, including seven books. He is currently research co-ordinator with the Federal Institute for Materials Research and Testing (BAM) in Berlin, Germany in the field of environmental engineering. Franz-Georg Simon studied chemistry at Frankfurt and Mainz Universities (diploma and Ph.D.), and spent 8 years in industry where he headed the research division of a major multinational company. He holds a Master's degree in Business Engineering, published some 40 papers, edited one book and is currently Head of the Division of Waste Treatment and Remedial Engineering within the Federal Institute for Materials Research and Testing (BAM). Douglas I. Stewart is a senior lecturer in Civil Engineering at the University of Leeds. He was previously a visiting assistant professor at Oregon State University (1989-1991), a research assistant at the University of Cambridge (1983-1988), and an assistant engineer working for Soil Mechanics Ltd (1982-1983). He has a B.Sc. in Civil Engineering from University College London (1982), and has an M. Phil. and a Ph.D. in Soil Mechanics from the University of Cambridge (1986 and 1990).

Acknowledgements The authors acknowledge the sponsoring of major parts of the research described in this book by the European Commission within the Thematic Programme "Energy, Environment and Sustainable Development" (EESD), Key Action "Sustainable Management and Quality of Water", of the Fifth Framework Programme for Research and Technological Development (project name: "Long-Term Performance of Permeable Reactive Barriers Used for the Remediation of Contaminated Groundwater", project acronym: PEREBAR, contract No: EVK1-CT-1999-00035).

Long-term Performance of Permeable Reactive Barriers K.E. Roehl, T. Meggyes, F.-G. Simon, D.I. Stewart, editors 9 2005 Elsevier B.V. All rights reserved.

Chapter 1 P e r m e a b l e reactive barriers Karl Ernst Roehl, Kurt Czurda, Tam~s Meggyes, Franz-Georg Simon and D.I. Stewart

A. Introduction The problem of anthropogenic groundwater contamination is widespread throughout Europe. Due to the large number of contaminated sites that require treatment, and in light of the incorporation of eastern European countries with their sometimes appalling ecological problems into the European Union, there is an urgent need for cost-effective risk management. In the field of contaminated groundwater, risk management typically involves remediation technologies for the control of the contaminant source and the management of contaminants along the pathway (CLARINET, 2002). The objective of the risk management is to break the link between pollutant source and receptor (such as drinking water resources) by managing or blocking the pathway. Groundwater remediation schemes are widely used to achieve this objective, mainly based on active methods such as pump-and-treat techniques. Recently, passive treatment methods have become more widely accepted as cost-effective and sustainable solutions to various types of water and soil pollution problems. These methods include: 9 natural attenuation, suitable primarily for the control of organic pollutants; 9 wetland systems, as used especially for the management of mine effluents; 9 permeable reactive barriers (PRBs) for groundwater remediation. This introductory chapter intends to give a brief overview of the application of PRBs to groundwater remediation. It is also the objective of the authors to encourage further reading by giving a selection of references covering the already quite extensive literature in the field of PRB research and application.

B. Concept of permeable reactive barriers Passive in situ groundwater remediation using PRBs is a relatively new and innovative technology with a high potential to significantly reduce the cost of treating contaminated shallow aquifers and therefore contribute to the preservation of groundwater resources. A PRB is a subsurface structure situated across the groundwater flow path downstream of a contaminant plume (Fig. 1. l). The barrier is constructed totally or in part from material that is hydraulically permeable and reacts with the passing groundwater to remove the

2

K.E. Roehl et al.

Figure 1.1. Schematicdepiction of the PRB concept (GW, groundwater flow direction).

contaminants from the groundwater. Processes taking place in the reactive material of the barrier include physical, chemical or biological contaminant retention reactions and the reactions of other groundwater constituents with the material. Suitable materials for use as reactive components in PRBs are elemental iron, activated carbon, zeolites, iron oxides/ oxyhydrates, phosphates, clay minerals and others. The most commonly used mechanisms are redox and sorption reactions. The choice of reactive materials and retention mechanisms are dependent on the type of contamination to be treated by the barrier system. The concept of PRBs was first developed in North America, with pioneering work conducted at the University of Waterloo in Canada. Initially the activities, including first pilot field tests, focussed on "funnel-and-gate" systems and the abiofic reductive dehalogenafion of chlorinates and recalcitrant compounds by elemental iron (Gillham and O'Hannesin, 1992,1994; Starr and Cherry, 1994; Tratnyek, 1996; Vidic and Pohland, 1996; Sivavec et al., 1997; Tratnyek et al., 2003). During the 1990s, research activities on PRBs increased significantly leading to a number of new approaches in terms of PRB design, suitable reactive materials and target contaminants. Amongst the first and most widely studied metal compounds treated by PRBs are chromate (Powell et al., 1995; Blowes et al., 1997) and uranyl (Cantrell et al., 1995; Bostick et al., 1997; Dwyer and Marozas, 1997) which are usually treated by reductive processes using, for example, elemental iron. The use of PRBs for groundwater protection or remediation has also been studied in other fields, e.g. the treatment of metals-containing mine waters (Morrison and Spangler, 1992, 1993; Thombre et al., 1997; Waybrant et al., 1998; Benner et al., 1999; Naftz et al., 1999; Younger, 2000; Waybrant et al., 2002). According to Blowes et al. (2000), the treatment of inorganic anions and cations can be grouped into abiotic reduction and immobilisation (mostly by elemental iron), biologically mediated reduction and immobilisation (bacterial sulphate reduction and precipitation of metals as sulphide), and adsorption and precipitation reactions. PRBs are defined by the US Environmental Protection Agency as "passive in-situ treatment zones of reactive material that degrades or immobilises contaminants as ground water flows through it. PRBs are installed as permanent, semi-permanent, or replaceable

Permeable reactive barriers

3

units across the flow path of a contaminant plume. Natural gradients transport contaminants through strategically placed treatment media. The media degrade, sorb, precipitate, or remove chlorinated solvents, metals, radionuclides, and other pollutants" (EPA, 1999). The substantial deviation from common remediation techniques is that the contaminant plume, and not its source, is treated (Schad and Grathwohl, 1998). The selection of the reactive material to be used in a PRB depends on the type of contaminant and the remediation approach (contaminant removal mechanism). In general, contaminants can be removed from polluted water by the following processes: Application of chemical or biological reactions that lead to the decomposition of contaminants and the formation of harmless compounds which are either retained in the barrier or released downstream. 9 Precipitation. Immobilisation of contaminants by formation of insoluble compounds (minerals), often after first reducing the contaminant to a less-soluble species. The immobilised contaminants remain in the barrier material. 9 Sorption. Immobilisation of contaminants by adsorption or complex formation. The immobilised contaminants remain in the barrier material. 9 Degradation.

Frequently, groundwater treatment can involve a combination of these processes which cannot be individually distinguished. Nowadays the most widely used approaches for PRBs can be grouped into two categories: reductive barriers and sorption barriers. Reductive barriers employ mechanisms that lead to the reduction of the target compound, or parts thereof, to achieve decomposition or immobilisation of that compound. Barriers utilising surface reactions that lead to immobilisation of the target contaminants by adsorption, ion exchange, co-precipitation, solid-solution formation, etc. without altering the chemical state of the contaminant are usually termed as sorption barriers. To date, two main types of PRB have been used in the field. These are (i) continuous reactive barriers enabling a flow through its full cross-section, and (ii) "funnel-and-gate" systems (Starr and Cherry, 1994) in which only special "gates" are permeable to the contaminated groundwater. The continuous PRB configuration is characterised by a single reactive zone installed across the contaminant plume, while the "funnel-and-gate" system consists of an impermeable wall that directs the contaminated plume through one or more permeable gates within the wall (Fig. 1.2). The choice between the two configurations depends on the hydrogeological characteristics of the site, the technical applicability of the barrier placement, and on

Figure 1.2. Basic types of PRB configuration: (a) continuous barrier, (b) "funnel-and-gate" system.

4

K.E. Roehl et al.

the cost of the reactive material. When a high-cost reactive material is used, the "funneland-gate" configuration is preferable since the reactive zone requires less material. If a cheap material can be used, it is more profitable to avoid the construction of the impermeable side-walls by employing a continuous barrier. New approaches to the PRB concept can be imaginable which modify the initial ideas. Containment of a contaminated site can be coupled with "gates" comprising reactors treating contaminants leached from the soil by infiltrating rain water. Contaminated surface and ground water from polluted sites could be collected in trenches or drains and treated in underground reactors before being discharged into a nearby river or sewer. Another option using in situ reactive zones is the so-called Geosyphon system, which utilises gravitational hydraulic gradients in pipes to draw contaminated groundwater through a treatment reactor filled with suitable reactive material (Jones et al., 2002). For greater depths, where conventional PRBs cannot be easily constructed, Freethey et al. (2002) have proposed the deep aquifer remediation tool (DART) which consists of nonpumping wells filled with reactive material. These examples show that a variety of approaches and solutions exist that make it possible to adapt the PRB technique to specific site conditions and contamination situations. As PRBs are low maintenance, subsurface installations, they are suitable for groundwater remediation schemes not only in typical industrial areas but also in residential and typical urban areas such as shopping centres, parking lots, industrial parks, etc. In all cases, a number of factors have to be considered and addressed during the planning and installation of a PRB system, e.g.: 9 the property boundaries; 9 the position of underground utilities such as water and sewage pipes, gas lines, cables, etc.; 9 any disruption of site activities that may be caused by the construction of the PRB; 9 the need to dewater the excavation pit, and the disposal of potentially contaminated water and soil; 9 the logistics and on-site management of material placement (e.g. quality control, homogenous filling of the reactors, prevention of dust, etc.); 9 human health and safety issues; 9 unforeseen ground conditions (particularly undetected underground utilities and abandoned foundations from demolished structures, etc.). Furthermore, the characteristics of the contaminated site need to be fully investigated when planning a PRB system. Crucial parameters include the hydraulic setting, the types and concentrations of contaminants, the total mass of contaminants, and the groundwater composition. Feasibility studies are nearly always necessary, usually involving the following steps: 9 choice of suitable remediation approach (contaminant removal mechanism) and reactive material; 9 column experiments (and other experiments quantitatively measuring the contaminant attenuation capability of the reactive material); 9 estimation of required residence time; 9 calculation of reactive zone thickness.


5

In PRBs, the residence time of the contaminant in the reactive material must be long enough to allow a decrease of the contaminant concentrations down to an acceptable level (the remediation target). For a given contaminant and reactive material the required residence time is a function of the reaction rate and the equilibrium constant. As a PRB is basically a flow-through cell there is a continual re-establishment of equilibrium (or at least the tendency towards re-establishing equilibrium) as the groundwater passes through the barrier. For example, a given reactive material could reach equilibrium very rapidly with a contaminant, but if the initial concentration is high and the equilibrium constant is low, then a long flow path may be required to reduce the contaminant concentration to an acceptable level. Alternatively, a slow reacting material with a high equilibrium constant may reach an acceptable exit concentration in a relatively short residence time, without ever reaching equilibrium. As a range of factors effect reaction rate, the retention time required to treat the groundwater at a particular site with a specific reactive material should always be determined in a feasibility study (e.g. by column experiments). Following these considerations, the minimum thickness of the reactive barrier can be calculated from the groundwater flow velocity in the barrier and the required residence time (Gavaskar et al., 1998; Carey et al., 2002): b = v X tR X SF

(1.1)

where b is the barrier thickness (m), v the flow velocity in the barrier (m/s), tR the residence time (s) required and SF a safety factor.

C. Reactive materials for contaminant attenuation Reviews on reactive materials suitable for use in PRBs for the removal of inorganic and organic compounds from groundwater are available in a number of publications (Rael et al., 1995; Baker et al., 1998; Gavaskar et al., 1998; Scherer et al., 2000; Simon and Meggyes, 2000; Roehl et al., 2001; Xenidis et al., 2002), also with a focus on mine waters (Blowes et al., 2000; Younger, 2000; Wolkersdorfer and Younger, 2002), and do not need to be repeated here. Thus, the following remarks are intended to summarise the most common techniques and point to emerging issues and innovative approaches. To date, the most widely used reactive material is granular elemental iron (zero-valent iron, Fe~ An extensive review on the iron technique has been given recently by Tratnyek et al. (2003). The widespread use of elemental iron is attributed to its ability to act as a strong reducing agent in groundwater causing abiotic reductive degradation of organic substances such as chlorinated hydrocarbons and some aromatics. Reductive dehalogenation results in the transformation of chlorinated hydrocarbons to less halogenated and finally halogen-free compounds (Tratnyek et al., 2003): Fe ~ + RCI + H + ~ Fe 2+ + RH + CI-

(1.2)

A typical pathway is the decomposition of tetrachloroethene (perchloroethene, PCE) via the intermediates trichloroethene, dichloroethene and chloroethene (vinylchloride) to ethene. By-products forming in this process are related to the corrosion of the elemental iron in the aqueous system and the consumption of acidity. Besides the release of chloride

6

K.E. Roehl et al.

ions, such by-products that might affect the functioning of a subsurface remediation system based on this process are the formation of hydrogen gas and the precipitation of secondary minerals such as hydroxides, carbonates and sulphide. Some attention has been paid recently also to the reductive immobilisation of inorganic compounds such as chromium, nickel, lead, uranium, sulphate, nitrate, phosphate, arsenic and molybdenum, among others. For example, highly soluble chromate - a known carcinogen - can be removed from groundwater in iron PRBs by a coupled reduction/precipitation mechanism (Blowes et al., 2000): 0 Fe~solid) + CrO 2- + 8H + ~ Fe3+ + Cr3+ + 4H2 O

( 1.3)

(1 - x)Fe 3+ + xCr 3+ + 4H20 ----}Fe(1_x)CrxOOH(solid) -k- 3H +

(1.4)

Another reductive mechanism, particularly important for the treatment of - often acidic - mine waters, is bacterial sulphate reduction as supported by organic materials such as compost, wood chips, sawdust, etc. (Benner et al., 1999; Blowes et al., 2000; Waybrant et al., 2002). In this approach, the reduction of sulphate to sulphide is utilised for the removal of metals from contaminated water by precipitation as sulphide (Blowes et al., 2000): 2CH20(solid ) + SO 2- + 2H + ~ H2S + 2C02 + H20

(1.5)

Me 2+ + H2S --~ MeS(solid) + 2H +

(1.6)

CH20 represents the organic carbon present in a suitable organic PRB filling. The production of alkalinity and resulting rise in pH increases the efficiency of the system by precipitation of metals as hydroxides. Due to the high efficiency of the amalgamation process in removing Hg 2+ from aqueous solution, the use of elemental copper shavings (Cu ~ for the removal of mercury from contaminated water is suggested by Huttenloch et al. (2003), employing a sequential system of mercury amalgamation followed by the removal of copper mobilised from the shavings by an ion exchanger such as zeolite. Hg 2+ is reduced to Hg ~ by Cu ~ forming a H g - C u amalgam, and stoichiometric amount of Cu 2+ is released: Hg 2+ + Cu ~ Hg + Cu 2+

(1.7)

Hg + Cu ~ CUHgAM

(1.8)

For Reaction (1.7), the Gibbs Energy can be calculated using the standard potentials from the electrochemical series (Cu ---, Cu 2+ + 2e, E ~ = -0.3402V; Hg 2+ + 2e ~ Hg, E ~ + 0 . 8 5 1 V ; resulting in a standard potential difference A E = O . 5 1 0 8 V ) as AG = - 9 8 . 5 7 kJ/mol. The resulting, clearly negative Gibbs Energy shows that the mercury reduction as described in Reaction (1.7) is the favoured reaction triggering the amalgamation. Elemental copper can be used in form of copper shavings that can be obtained as a recycling product (scrap metal). Although tested only at laboratory scale, the removal of mercury from water streams by amalgamation to copper shavings appears to be a promising approach due to the high Hg 2+ retardation coefficients and fast reaction kinetics that were achieved (Huttenloch et al., 2003).


7

Sorption barriers are PRBs utilising retention mechanisms that lead to a fixation of the target contaminants to the matrix of the reactive material (Roehl et al., 2001). In this context the term sorptive barrier usually describes any barrier where the removal process does not destroy or change the oxidation state of the contaminant. Processes can include surface adsorption, ion exchange, surface complexation, precipitation and coprecipitation. The manner and the strength of fixation is of great importance regarding the possibility of remobilisation, and is strongly influenced by geochemical parameters like the concentration, solubility and speciation of the contaminants and co-solvents, and the prevalent pH and Eh conditions. In sorption barriers, the most important retention mechanisms are adsorption, ion exchange and precipitation. The extent of adsorption is mainly governed by the size of the specific surface area of the sorbent. Ion exchange is governed by the cation exchange capacity (CEC) or the anion exchange capacity (AEC) of the sorbent, as appropriate. Precipitation processes may be used to immobilise heavy metals as, e.g. carbonates, phosphates, hydroxides or sulphide. The sorption processes leading to an immobilisation of contaminants can be summarised by the retardation coefficient R which can be deducted from the relation between sorbed and aqueous concentrations of the contaminant (Yong et al., 1992) R = 1 -~

Pd Of(c) n 0c

(1.9)

where Pd is bulk density and n porosity of the sorptive material, c the concentration of the contaminant in solution, and f(c) a linear or non-linear relationship between sorbed and aqueous concentrations of the contaminant (i.e. Of(c)/Oc is the slope of the sorption isotherm). A large number of materials that are able to sorb and trap certain contaminants, and therefore immobilise them from the groundwater, are imaginable. The efficiency of the immobilisation mechanisms in terms of its sorption capacity, selectivity, reaction kinetics and bonding strength is of great importance. The target contaminants have to be fixed to the reactive material in a way that they are not easily remobilised and subsequently released to the groundwater. The reactive materials also need to be available in a form that ensures a sufficiently high hydraulic permeability and exhibit a non-harmful behaviour towards the environment. The choice of reactive material is dependent on the type of contaminant to be treated by the barrier system. Possible materials for the use as reactive components in sorption barriers are activated carbon (Sontheimer et al., 1988; Schad and Grathwohl, 1998; Han et al., 2000; Tiehm et al., 2000), zeolites (e.g. Ouki et al., 1993; Cantrell et al., 1994; Pansini, 1996; Czurda, 1999; Anderson, 2000; Czurda and Haus, 2002; Park et al., 2002), iron oxides/oxyhydrates (Morrison and Spangler, 1992; Morrison et al., 1995a,b; Moyes et al., 2000), phosphate minerals (Ma et al., 1993; Xu and Schwartz, 1994; Fuller et al., 2003), and surface-modified minerals such as organophilic zeolites (Haggerty and Bowman, 1994; Cadena and Cazares, 1996; Bowman, 1999; Huttenloch et al., 2001), diatomites (Huttenloch et al., 2001) or clays (Smith and Jaffe, 1994; Smith and Galan, 1995; Lundie and McLeod, 1997; Zhu et al., 2000). A wide selection of low-cost sorbents, including mineralic and non-mineralic materials, such as bark, chitin, chitosan, lignin, seaweed and algae, xanthates, zeolites, clay, fly ash, peat, moss, etc., for the sorption of heavy metals is discussed and evaluated

8

K.E. Roehl et al.

by Bailey et al. (1999). The authors conclude from their literature review that for Pb, Cr, Cd and Hg the highest sorption capacities were found for chitosan, zeolites, lignin and seaweed. Activated carbon is to date the most widely used material in sorption barriers. The adsorption of organic compounds to activated carbon is a well-established method for on-site or off-site treatment of polluted water. In granular form, activated carbon appears to be highly suitable for the use in permeable barriers. Due to its large specific surface area (around 10OOm2/g by N2-BET is a typical value) and the presence of different types of surface functional groups, activated carbon shows a high adsorption capacity for many organic and inorganic contaminants (Sontheimer et al., 1988; Grathwohl and Peschik, 1997; Schad and Grathwohl, 1998; Han et al., 2000; Tiehm et al., 2000; Schad et al., 2001; Kraft and Grathwohl, 2003). Phosphate minerals such as hydroxyapatite and biogenic apatite (e.g. fishbone) enable the removal of metals from contaminated water by sorption and precipitation or a combination of both mechanisms, as described for lead (Ma et al., 1993; Xu and Schwartz, 1994; Admassu and Breese, 1999), antimony (Leyva et al., 2001) and uranium (Arey et al., 1999). Other materials bear some potential for use in special cases, such as organophilic zeolites and diatomites with silane surfaces (Huttenloch et al., 2001) and organo-clays (Ake et al., 2001). The removal of inorganic oxyanions such as sulphate, chromate and selenate from aqueous solution by clinoptilolite-rich zeolite modified by HDTMA was studied by Haggerty and Bowman (1994). The use of pillared clays in PRBs to remove PAHs and heavy metals from contaminated groundwater is discussed by Lundie and McLeod (1997). The choice of material should be based on the following criteria (after Gavaskar et al., 1998): 9 Reactivity. The required residence time is directly related to the reaction rate and

9

9

9

9

9

equilibrium constant of the contaminant with the reactive material, and should be quantitatively evaluated. It is desirable to have low residence times and thus high reaction rates in order to keep the barrier' s thickness within acceptable limits. Stability. The material will be expected to remain active for long periods of time because its replacement may not be easily achieved. Stability upon changes of pH, temperature, pressure and antagonistic factors is also required. Availability and cost. The amount of reactive material required for the construction of a reactive barrier may be large and therefore it is essential to have considerable quantities available at low prices. Hydraulic performance. The hydraulic conductivity (permeability) of the reactive material depends on its particle size distribution, and must be greater or equal to that of the surrounding soil, so an optimum particle size that would provide appropriate permeability must be determined. Environmental compatibility. It is important that the reactive media does not form any unwanted by-products with the contaminant and does not dissolve or release any undesirable substance and thereby become a new source of contamination. Safety. Handling of the material should not generate any significant risk to workers' health.


9

D. Application and long-term performance of PRBs Permeable reactive barrier technology appears to be a promising approach to effective groundwater remediation even in complex cases where traditional "pump-and-treat" methods and/or microbiological techniques have proved unsuccessful (e.g. heavy metals being slowly leached from a contamination source; PAH with low bio-availability; contamination of heterogeneous sediments). Although the use of PRBs is limited to certain site conditions, where they are feasible they appear to be a good choice with good acceptance by end-users, especially in urban environment and built-up areas. Reasons for this include little land use, low visibility, no need for ancillary equipment such as containers, water tanks, pumps, that can impact on the landscape, and no noise nuisance from machinery, etc. Only a small number of pilot-scale and full-scale installations exist today, mainly located in the United States (Morrison, 1998; EPA, 1999) and some European countries, including Germany (Birke et al., 2003), and so practical experience with such systems is limited. More than 50% of the existing projects are based on the elemental iron technology (EPA, 2002). In Germany, l i pilot-scale and full-scale PRBs exist or are in the process of being implemented. Of these projects, seven are based on elemental iron, sometimes with supplemental techniques, three on granular activated carbon (GAC) and one on palladiummodified zeolites. A good deal of information on the PRB sites in Germany can be found on the website of the German Permeable Reactive Barrier Network "RUBIN" which also features an extensive English language section (http://www.rubin-online.de/). With a few exceptions (e.g. Ebert et al., 1999; Klein and Schad, 2000; Birke et al., 2003; Ebert et al., 2003) there is currently little published information available on the functioning and success of the German PRB systems. When PRBs are used as an alternative to conventional remediation methods, it is currently unclear how they should be regulated as little is known about the long-term behaviour of such systems. Therefore, as a relatively new approach with only little information available about its long-term performance, PRBs are currently not well accepted in Europe, although further developments are now being advanced by a number of research groups and institutions. However, in addition to the development of new barrier technologies (such as new reactive materials and barrier construction methods) more information is especially needed on the long-term behaviour of the reactive barriers and on the any processes that might affect the long-term success of the barrier. Some experiences with existing PRB installations are discussed in the following sections.

1. Elemental iron barriers In all of the pilot and commercial installations to date there has been little data collected on the long-term performance with respect to the build-up of surface precipitates or biofouling. The analysis of the formation rate of surface precipitates is critical to understand how long PRB installations will remain effective. Different types of surface coatings have been observed in different geochemical conditions which are determined by the composition of the reaction zone of the PRBs and the aquifer chemistry

10

K.E. Roehl et al.

(Puls, 1999; Klausen et al., 2003). The lifetime of a PRB might be limited by the reaction processes of the contaminated groundwater with the reactive material. Most information concerning long-term performance of PRBs is available for elemental iron (Fe o) barriers since iron is the most commonly used reactive material in PRBs. The corrosion process induced by water or dissolved oxygen in the influent groundwater reacting with the Fe ~ material leads to the formation of Fe 2+, hydrogen gas and OH-. The release of OH- causes an increasing pH within the reactive barrier. The values observed in laboratory and field studies are ranging from pH 8-11 (Gillham and O'Hannesin, 1994; O'Hannesin and Gillham, 1998; McMahon et al., 1999; Puls et al., 1999; Vogan et al., 1999; Klein and Schad, 2000; Ebert et al., 2003). At these conditions, iron salts are not soluble and precipitate as ferric oxyhydroxides (ferrihydrite), goethite (FeOOH), amakinite (Fe(OH)2) or magnetite (Fe304) (Gillham and O'Hannesin, 1994; Mackenzie et al., 1999; Phillips et al., 2000; Gillham et al., 2001; Morrison et al., 2001; Furukawa et al., 2002). The corrosion processes and the subsequent precipitation of minerals may lead to cementation and a decreasing permeability of the reactive material (Mackenzie et al., 1997; Phillips et al., 2000). Decreases in the reactivity of the iron media may also occur due to the formation of surface coatings. The longevity of the iron-material is also strongly influenced by the presence of cosolutes in the site groundwater. The long-term effects of dissolved carbonate, sulphate, silica, nitrate, chloride and natural organic matter (NOM) on the reactivity of Fe ~ was investigated in column studies by several authors. Bicarbonate, sulphate and chloride containing water have a significant effect on the corrosion rates of Fe ~ fillings resulting in the formation of mineral precipitations (Reardon, 1995; Dahmke et al., 1997; Gu et al., 1999; Klausen et al., 2003). Major mineral precipitations are siderite, calcium carbonates, iron oxides, iron oxyhydroxides, iron sulphide and sulphate or carbonate bearing green rust (Gu et al., 1999; Yabusaki et al., 2001; Furukawa et al., 2002; Klausen et al., 2003). Klausen et al. (2003) observed that high carbonate concentrations initially enhance the reactivity of Fe ~ material towards the dehalogenation of organic compounds, but the reactivity decreases in the long term. Krber et al. (2002b) attribute the decreasing dehalogenation rates of Fe ~ to precipitation of carbonates and the development of hydrogen gas by anaerobic corrosion. The H2 gas generated by iron corrosion supports the growth of microbial populations, increasing nitrate and sulphate reduction processes (Till et al., 1998; Gu et al., 1999; Krber et al., 2002b). The formation of H2 gas bubbles can reduce the porosity of the reactive material resulting in decreasing hydraulic permeability (Bokermann et al., 2000; Gilham et al., 2001). Microorganisms can reduce the porosity by forming a biofilm (Gu et al., 1999) or conversely the hydraulic permeability can also be increased by microorganisms consuming the H2 gas produced by the iron corrosion process (Scherer et al., 2000). Compounds like nitrate, silica or NOM have a passivating effect, reducing not only the iron corrosion rates but also the iron reactivity (Farrell et al., 2000; Schlicker et al., 2000; Gilham et al., 2001; Klausen et al., 2003). In the presence of nitrate the dehalogenation of chlorinated compounds is decreased indicating a competitive process between dechlorination and nitrate reduction (Farrell et al., 2000; Schlicker et al., 2000). Silica precipitates on the iron surfaces to form a silica film that hinders the contaminants reaching the active sites on the iron material (Klausen et al., 2003). NOM can be adsorbed


11

by iron(hydr)oxides which are present on the iron surface causing a decreasing reactivity of the iron surface (Klausen et al., 2003). These effects have been documented in laboratory analyses, usually based on the performance of column experiments, and field observations at a number of sites where full-scale PRB systems have been installed. Some experiences from such sites are summarised in the following. An elemental iron PRB to remove chromium and chlorinated solvents was installed in Elisabeth City, North Carolina. Mineral precipitates like calcium carbonate, iron hydroxy carbonates, carbonate green rust, hydrous ferric hydroxides, ferric oxyhydroxides and also iron mono-sulphide (mackinarite and greigite) were detected after 4 years operation time (Puls et al., 1999; Furukawa et al., 2002; Wilkin et al., 2002; EPA, 2004). The corrosion layer is greatest within the first 5 cm and decreases significantly within 20 cm from the up-gradient aquifer/iron interface (Puls et al., 2000). After 4 years of mineral precipitation and accumulation, a consistent surface coverage of the reactive material ranging in thickness from about 10-50 Ixm near the up-gradient interface to 5 txm at the down-gradient interface was found (Furukawa et al., 2002; EPA, 2004). During operation the average rates of inorganic carbon and sulphur accumulation are 0.09 and 0.02 kg/m 2 y, respectively, with local groundwater containing less than 400 mg/l of total dissolved solids (TDS) (Wilkin et al., 2002). The total porosity loss by mineral precipitation in the up-gradient first 2.5 cm within the iron media was found to be 5.9% maximum, and values decrease to less than 0.1% at distances greater than 8 cm from the up-gradient interface (EPA, 2004). Microbial activity is also observed at the site. After 4 years, the biomass accumulation ranged about 5-875 pmol of phospholipid fatty acids (PLFA) per g iron or between 1.02 • 105 and 1.78 • 107 cells per g iron matrix with the highest concentration measured at the up-gradient interface. The biomass consists of anaerobic sulphate and metal reducing bacteria (EPA, 2004). It is interesting to note that the hydraulic permeability and the reactivity of the Fe ~ material were not affected by the mineral precipitation or bioaccumulation during operation time (Wilkin et al., 2002). An Fe~ funnel-and-gate system with four gates was implemented at the Denver Federal Center, Colorado, for the treatment of volatile organic compounds. The system appears to work successfully during its operation time of 4 years except for one gate (gate 2) where a contaminant breakthrough was observed soon after the system was installed (Wilkin et al., 2002; EPA, 2004). The up-gradient groundwater contains 1000-1200 mg/l of TDS. The rates of inorganic carbon and sulphur accumulation are 2.16 and 0.80 kg/m2y, respectively (Wilkin et al., 2002). A greater build-up of mineral precipitates (predominantly carbonate and siderite) and biomass was identified in gate 2 (McMahon et al., 1999; EPA, 2004). The concentrations of inorganic carbon and sulphur in this gate are increased by about four times of the maximum amounts observed in the other gates. Also the biomass accumulation of 4100 pmol of PLFA per g dried iron or between 8.36 • 107 cells per g iron matrix is relatively high, compared to, e.g. the Elisabeth City site. The porosity loss of 14.2% at 2.5 cm into the iron media (up-gradient interface) after 4 years in gate 2 compared to 6% in gate 1 can be related to a decreasing reactivity of the iron media (EPA, 2004). The performance of a pilot-scale reactive barrier at an industrial facility in New York was investigated over a 2-year period (Vogan et al., 1999). During operation the total

12

K.E. Roehl et al.

porosity loss caused by carbonate precipitation was 10% (calcite, aragonite and siderite). The maximum CaCO3 content of 6% was found near the up-gradient interface, rapidly declining to values of less than 1% at a distance of 15 cm. Calculations based on declining carbonate content of the groundwater passing the barrier showed that 15 kg CaCO3 would have been deposited in a 1 m 2 • 0.46 m section of the iron barrier within 2 years. Collected core samples confirmed these assumptions. The system performance was not affected by the mineral precipitation or microbial growth. Based on these results a fullscale continuous barrier was installed. Kiilerich et al. (2000) calculated the amount of mineral precipitation in an iron barrier in Copenhagen, Denmark, based on changes of the dissolved inorganic species of the upgradient and the down-gradient groundwater. The precipitations are calculated to be 2.7 kg CaCO3, 2.7 kg FeCO3 and 0.8 kg FeS per ton iron and year. The precipitation of Fe(OH)2 that will take place due to corrosion of the iron was estimated by the hydrogen production and the release of Fe 2+ to values of 13.3 kg iron hydroxide per ton iron and year. No significant effect on the dehalogenation capacity of the iron barrier and no impact on hydraulic permeability was observed during the operation period of 15 months. Funnel-and-gate systems based on Fe ~ were installed both at the NAS Moffett Field site (April 1996) and at the Lowry AFB site (December 1995) to treat chlorinated solvents (primary contaminant: TCE). The up-gradient groundwater contains 820 mg/l TDS (Ca 180 mg/l, $04 360 mg/l, alkalinity 390 mg/l) and 1700 ~g/l TCE at the Moffett Field site and 2900 mg/l TDS (Ca 290 mg/l, $04 lOOOmg/l, alkalinity 530 mg/l) and 71 p~g/l TCE at the Lowry AFB site (EPA, 2004). At the Moffett site minerals like ferrous (hydr)oxide, siderite, aragonite, calcite, brucite and iron sulphide (mackinawite) were precipitated in the first few centimetres of the iron cell causing a porosity loss of 3% per year (Yabusaki et al., 2001; EPA, 2004). Decreasing sulphate concentrations in the iron material may result in precipitation of sulphate-containing green rust ([Fe4Fe2(OH)12][SO4 • 3H20]) or iron sulphide caused by the reduction of sulphate to sulphide. No microbial activity is observed in the remediation system. Similar mineral precipitations are observed at the Lowry AFB site, with a greater amount of carbonates deposited in the up-gradient portion of the barrier. Core samples collected near the upgradient interface contained 4 g CaCO3 per lOO g iron after 18 months of operation time (EPA, 2004). At the Y-12 pathway site in Oak Rich, Tennesse, an Fe ~ barrier was installed to remove uranium from groundwater. In the up-gradient Fe ~ material significant cementation was observed, caused by corrosion due to high nitrate levels in the local groundwater ranging from 38 to 822 mg/l (Phillips et al., 2000; EPA, 2004). The corrosion appears as finds of FeOOH ranging in thickness from 10-150 ~m, sometimes binding the iron fillings together (Phillips et al., 2000). Column tests conducted by Mackenzie et al. (1999) showed that the Fe ~ fillings were clogged with FeOOH as a hardened solid mass that can decrease hydraulic permeability. The corrosion of Fe ~ at the Y-12 site appeared to have degenerated 15-30% of the Fe ~ fillings in the cemented samples. Phillips et al. (2000) estimate that these Fe ~ fillings will be corroded in 5 - 1 0 years under the site specific geochemical conditions if corrosion continues. Other mineral precipitates observed at the site after 15 months of operation time are calcium carbonate, siderite, green rust, goethite and iron sulphide (Phillips et al., 2000; EPA, 2004). The accumulation of carbonates may


13

eventually prevent the corrosion of the iron fillings by limiting mass transfer to the iron surface, but in consequence the surface reactivity will decrease. The Monticello PRB was installed to treat uranium- and vanadium-contaminated groundwater. Column tests showed that Fe ~ is able to decrease these contaminants in the local groundwater to non-detectable levels. After percolation of 3000 pore volumes the iron material in the column experiments was less efficient due to corrosion products and accumulation of mineral precipitates (Morrison et al., 2001). After 2.7 years of operation of the full-scale PRB 8.8 ton of calcium carbonate and 24 kg of combined U and V beating minerals were precipitated (the local groundwater contains up to 295 mg/l calcium, 1180 mg/l sulphate, l 18 mg/l nitrate, 173 mg/l chloride, 430 mg/l alkalinity; EPA, 2004). More than 99% of the U and V precipitates are deposited in the up-gradient gravel/iron zone. The calcium carbonate minerals are distributed both in the gravel/iron and pure iron zones. The distribution of these minerals suggests that the precipitation reactions for U and V are rapid relative to those for Ca. The porosity loss during operation time is calculated to be 9.3% in the up-gradient gravel/iron zone and 3.2% within the iron zone (Morrison, 2003). A decreasing concentration gradient of Ca in the up-gradient gravel/iron zone shows that this zone lost some reactivity during its performance period of 2.7 years. Increasing U concentrations (from 0.2 to 185 Ixg/l) exiting the gravel/iron zone also indicate a loss of reactivity. Morrison (2003) calculated that the loss of reactivity of the iron fillings may have a greater impact to the long-term performance of the PRB than porosity reduction. A new design to extend lifetime of PRBs is suggested by Morrison (2003) for contaminants such as U which are removed from solution faster than calcium carbonate. The strategy is to decrease the residence time of the contaminated groundwater in the PRB. This can be achieved with perforated distribution pipes parallel to and spanning the length-of the barrier. At initial operation time the contaminated groundwater is guided by valves to distribution pipes near the down-gradient interface. As the reactivity at the down-gradient zones decreases, more up-gradient zones in the PRB can be activated. The shorter residence time may result in a smaller increase in pH values and decreasing carbonate mineral precipitation. A full-scale continuous Fe ~ barrier for the dehalogenation of chlorinated compounds (main contaminants: PCE and TCE) was installed in Rheine, Germany, in June 1998. The reactive barrier contains two different sections filled with (a) an iron sponge (IS) and (b) granular iron (GI) mixed with gravel. The results from preliminary column tests on iron sponge in combination with thermodynamic modelling are used to predict the long-term performance of the Fe ~ barrier. The prediction of the changes in groundwater chemistry caused by biogeochemical reactions as well as the degradation efficacy of > 9 9 % (concerning the iron sponge section) show very good conformity with the 5-year monitoring data of the PRB (Ebert et al., 2003). The monitoring results indicate that the prognosis on long-term performance of 5 - 6 years based on the column tests would be exceeded by the Pdaeine PRB. However, a decreasing dehalogenation rate from initially 99 to 75-90% (since May 2001) has been observed in the granular iron section. In addition to the conventional monitoring program tracer tests are performed in order to identify the problems limiting the performance of this section of the PRB (Parbs et al., 2003). Conservative and reactive tracers proved to be useful tools to determine the hydraulic permeability and reactivity of the Fe ~ filling throughout the barrier. A presupposition for a suitable reactive tracer is the specific reactivity solely with iron. A good

14

K.E. Roehl et al.

indicator for the current reactivity of the iron filling are, e.g. natural components of the groundwater like nitrate or TIC. Increasing concentrations of these compounds over time in the down-gradient groundwater indicate a decreasing reactivity of the barrier material. The application of nitrate (reactive tracer) and lithium bromide (conservative tracer) at the Rheine site show that the loss in reactivity of the granular iron/gravel zone might be caused by the formation of preferential flow paths due to lack of homogenous dispersion of the Fe ~ and gravel during construction. Similar tracer tests are performed at the BEKA site in Ttibingen, Germany, where a full-scale Fe ~ funnel-and-gate system consisting of three gates was installed in 1998 for the treatment of chlorinated hydrocarbons (Klein and Schad, 2000). Initially, the contaminants were removed to concentrations below I O Ixg/l. Six months after installation a slight increase in the contaminant concentrations (but still < 10 ixg/l) was observed down-gradient of gate l. Within the iron zone no chlorinated hydrocarbons could be detected. Also the common geochemical changes in groundwater composition within the gate, as caused by reaction with Fe o, were observed. The results indicated that the iron zone successfully dehalogenates the contaminants (Klein and Schad, 2000). The tracer tests show that the formation of preferential towpaths is obviously responsible for slightly increased concentrations of chlorinated hydrocarbons down-gradient of the treatment system (Parbs et al., 2003). These towpaths are probably caused by problems during the construction of the funnel-and-gate system or by inhomogeneous settling of the reactive material due to the great differences in density between the Fe ~ fillings compared to the aquifer material and the slurry walls, respectively. The continuous barrier at the Canadian Forces Base in Borden, Ontario shows an excellent performance since the time of installation in April 1996 (O'Hannesin and Gillham, 1998). The PRB consists of 22% of granular iron and 78% of sand in order to treat chlorinated organic compounds. Approximately 90% of the contaminants are removed by dehalogenation within the barrier and no decrease in performance was observed during 5 years of operation. The contaminated groundwater contains 278 mg/l calcium, 609 mg/l sulphate and the alkalinity is 169 mg/l. After 2 years of operation there was no evidence for precipitation and cementation in the collected core samples. The levels of biological activity were low. Little precipitation of calcium and iron carbonates was found within the first few millimetres of the barrier at the upgradient interface of the PRB after 4 years of installation (O'Hannesin and Gillham, 1998). The Sunnyvale Fe ~ PRB was installed in 1994 to treat dissolved chlorinated hydrocarbons (Sorel et al., 2003). The results of more than 5 years of monitoring show that the contaminants are dehalogenated below groundwater cleanup standards within the iron zone. Mineral precipitates are observed in the up-gradient pea gravel section protecting the iron filling from significant porosity loss. The reactivity of the iron is not affected by mineral precipitation or biological activity during operation time. Measurements of the hydrogen concentrations (generated by iron corrosion process) are performed, that may be an indicator for evaluation of the long-term performance of a PRB. The background hydrogen concentrations in the aquifer is about 2-15 nM. Within the iron zone the hydrogen concentrations increased up to the solubility limit of greater than 6OO,OO0nM. The hydrogen concentrations decreased to values of 1OOnM I0 ft up-gradient of the PRB indicating a presence of a hydrogen "halo" around the iron zone. The high concentration of dissolved hydrogen within the barrier shows that the


15

corrosion process remains strong even 6 years after installation of the reactive barrier (Sorel et al., 2003). An 18-month performance monitoring of a pilot-scale funnel-and-gate barrier at the Somersworth site in New Hampshire was conducted in order to collect data that could improve a full-scale PRB design (Sivavec, 2000; Sivavec et al., 2003). Extensive biodegradation occurs in the upgradient aquifer, so 50% of the chlorinated hydrocarbons were reduced by microbial activity. Within the barrier the contaminants were dehalogenated by the iron filling to non-detectable levels. The microbial growth within the barrier is not greater than observed in the surrounding aquifer. The up-gradient groundwater contains approximately 400 mg/l of TDS. Mineral precipitates (calcium and ferric carbonates) are identified with highest values at the up-gradient iron/pea gravel interface and decreases to background levels within the first 6 in of the iron zone. The porosity loss of 3% in the interface during operation time is less than predicted by laboratory studies. At the Area 5, Dover Air Force, a Fe ~ funnel-and-gate system (two gates) was installed in December 1997 to remediate dissolved chlorinated solvents (Yoon et al., 2000). During the monitoring time of 2 years the contaminants are reduced below target levels. In each gate a pre-treatment zone is placed to remove dissolved oxygen in the influent groundwater in order to extend the lifetime of the reactive iron material. Both types of pre-treatment media (gate l: 10% iron and sand, gate 2: 10% pyrite and sand) are able to remove the dissolved oxygen before entering into the reactive cells. As a result only a small amount of precipitation and corrosion built-up on the iron surfaces is observed in both cells since installation in the low-alkalinity aquifer. A pilot-scale iron barrier was installed at the south of the city of Sydney, Australia in February 1999 to treat chlorinated solvents (Duran et al., 2000). The degradation of the volatile compounds ranged between 81 and 96%. After 9 months, a decline in the degradation rate of PCE is observed in zones where the dissolved organic carbon (DOC) concentration was the lowest. This is contrary to the results of the column studies (which indicate an inhibition of degradation rates when DOC was introduced into the column) and will be further investigated. The up-gradient groundwater contains 1610 mg/l of TDS, but due to the low pH of 4.8, only 46 mg/l of bicarbonate was found. The sulphate (53 mg/l) and the sulphide (34 mg/l) contents of the up-gradient groundwater precipitate as iron sulphide. The loss of porosity caused by iron sulphide is estimated to approximately 1.3 % per year and would not affect the barrier performance. The iron barrier at the Almeda Naval Air Station (near San Francisco) was installed to treat organic solvents (Devlin et al., 2000). With a removal of > 95% the barrier does not completely degrade the entering organic compounds (cDCE up to 200 mg/l and VC 40 mg/l). Prior to installing the barrier, preliminary column studies with groundwater from the Almeda site (with high TDS concentrations ranging from 800 to 8700 mg/l) were conducted. The cDCE concentration of 35 mg/l was significantly below the maximum concentration in the heterogeneous plume. The results showed that a 37-cm thick barrier would be sufficient for treating the contaminants. The final barrier was installed with a conservative path length of 1.5 m to remediate high contaminant concentrations according to calculations with the pseudo-first-order model. Due to the inefficient treatment of the iron barrier further column studies were performed to investigate surface saturation effects depending on input concentration (20-300mg/1). The studies showed that the

16

K.E. Roehl et al.

breakthrough of the contaminants was a result of slower reaction kinetics at high concentrations due to surface saturation effects. Consequently, the inappropriate use of the pseudo-first-order model can result in significantly undersized barriers if the contaminant concentration in the plume exceeds those used in the preliminary studies. As shown in the case studies described above a typical problem of iron barriers is the formation of corrosion products and other precipitates on the iron surfaces which might cause a loss in reactivity over time. The lifetime of a barrier could be extended if the coatings could be removed from the iron surfaces. Clausen et al. (2000) investigated the application of ultrasonic energy to regenerate iron barriers with the goal of enhancing or restoring the rate of TCE degradation. Laboratory and field studies were performed to examine the impact of ultrasound on iron under different conditions. The degradation of TCE on iron is dependent on the surface area. Consequently, reaction rates decrease with increasing surface coatings on the iron media. The results of the experiments showed that the ultrasound technique is able to remove corrosion products and precipitates resulting in restoring a fresh iron surface. A sonication period of 30 min significantly improves the first order constant for TCE degradation. The long-term performance of PRBs is strongly dependent on the groundwater chemistry, groundwater flow rates and contaminant concentrations at the remediation site. The observed mineral phases at different PRB sites are controlled by the groundwater constituents, which naturally vary depending on the biogeochemical setting of the sites. Groundwater with high TDS has a greater potential for mineral precipitation, and especially high nitrate and sulphate concentrations accelerate iron corrosion. The overall hydraulic performance of a reactive barrier can be affected when the precipitates are not uniformly distributed in the reactive material resulting in preferential flowpaths and decreasing reactivity (Kamolpornwijit et al., 2003). At most Fe ~ PRB installations the main mineral and biomass build-up is observed in the first few centimetres of the barrier causing a porosity loss ranging from 1 to 5% per year. Other factors are gas production, microbial activity and transport of colloids within the barrier (Eykholt et al., 1999). These processes affect the residence time of the contaminated groundwater in the barrier resulting in higher contaminant concentrations at the effluent interface. However, at this time at most sites there is no indication that the mineral precipitates are affecting the PRB efficacy, so most barriers will function for at least 1O- 15 years before rejuvenation or iron replacement will be necessary (Vogan, 2003). Elder et al. (2002) investigated by numerical models how aquifer and barrier heterogeneity affect influent and effluent concentrations for PRBs. To be able to balance the heterogeneity a safety factor in barrier design is suggested by the authors. Benson (2003) used numerical models to simulate flow, transport and geochemical reactions in a heterogeneous aquifer to examine mineral precipitation and porosity reduction in PRBs under various conditions. The results indicate that the performance problems like porosity loss and impact on hydraulic conductivity become significant only until the PRB is at least 25 years old. 2. Other reactive materials

The use of organic carbon rich materials to promote sulphate reduction processes has already been described by Reaction 1.5 and 1.6. A full-scale continuous barrier was installed in August 1995 at the Nickel Rim Mine Site in Ontario, Canada, to induce


17

bacterially mediated sulphate reduction and subsequent metal sulphide precipitation (Benner et al., 1997). The contaminant plume at this site is characterised by elevated concentrations of sulphate (2400-4800 mg/l), Fe(II) (250-1300 mg/l) and Ni (10 mg/l) and a pH between 4 and 6. The reactive mixture contains municipal compost (40%), leaf compost (40%) and wood chips (20%). Pea gravel was added to maintain hydraulic conductivity. Nine months after installation the sulphate concentration decreased to 200-3600 mg/l, Fe(II) decreased to 1-40 mg/l (removal of >90%) and Ni to values < 0.1 mg/l, respectively. In addition, the pH increased to value around seven across the barrier and the alkalinity (measured as CaCO3) increased from 0 - 6 6 to 690-2300 mg/l. The PRB converted the aquifer from acid-producing to acid-consuming. The populations of sulphate-reducing bacteria were lO,OOOtimes greater and the microbial activity was l0 times higher within the reactive mixture compared to the up-gradient aquifer. The concentration of dissolved sulphide was increased by 0.2-120 mg/l resulting in major precipitation of iron mono-sulphide (mackinawite) (Benner et al., 1999). The iron and sulphate removal occurred at a 1:1 molar ratio limiting the removal of sulphate, because when all Fe(II) is precipitated as iron sulphide approximately half of the sulphate concentration remains in solution. Other mineral precipitates such as siderite and gypsum, but also formation of organo-sulphide compounds were observed within the barrier. The accumulation of precipitates caused a loss in porosity of 1% within 3 years and no changes in hydraulic conductivity was observed. After a 3-year monitoring period, the rate of sulphate removal within the barrier decreased by 30% from an initial rate of 58-40 mmol per litre and year. Over the same time the rate of Fe(II) removal declined by 50% from 38 to 18 mmol per litre and year (Benner et al., 2002). The organic matter used in the barrier contained a high chloride concentration leaching to the down-gradient aquifer (background aquifer concentration: 7 mg/l, concentration down-gradient of the barrier: up to 6000 mg/l). The transport of chloride indicates that the flow through the barrier was heterogeneous with higher flow velocities in the central section of the barrier compared to the top and the bottom. There is a positive correlation between increased concentrations of iron and sulphate and lower alkalinity down-gradient of zones of higher hydraulic conductivity, which was probably the result of reduced residence time. As suggested by modelling the velocities through the middle of the barrier were approximately three times faster (residence time about 60 days) than at the top and base (residence time about 165 days). The removal efficiency is also dependent on temporal variations resulting in a decreased organic carbon availability and reactivity over time and seasonal variations in the rate of sulphate reduction. The temperatures in the aquifer ranged from 2~ in winter (low reduction rate) up to 16~ in summer (the reduction rate was doubled compared to winter). However, 3 years after barrier installation >1000 mg/l sulphate and >250 mg/l iron are still being removed from the contaminated groundwater, showing a sufficient long-term performance of the sulphate-reducing barrier. The long-term performance of four pilot-scale sites treating nitrate contaminated wastewater is described by Robertson et al. (2000). The monitoring was performed through 6 - 7 years from 1992 to 1999. Nitrate can be remediated by an organic carbon source under anaerobic conditions due to heterotrophic denitrification, whereby nitrate is biodegraded to N 2 gas: 5CH20 + 4N03 --~ 5C02 + 2N 2 + 3H20 -k- 4 0 H -

(1.1O)

18

K.E. Roehl et al.

At two sites, Killarney and Borden in Ontario, Canada, reactive barriers were installed as horizontal layers underneath a septic system infiltration bed. Both reactive barriers consist of 15 vol.% waste cellulose material. The Killarney barrier is subdivided into three layers: the uppermost containing sawdust, the middle layer containing leaf compost and the bottom layer containing unprocessed grain seed. The seasonal average loading rate was 61/day until 1993 and increased up to 13 l/day resulting in decreasing residence time of the nitrate contaminated water in the barrier from 40 to 17 days. The average influent nitrate-N concentration is 57 mg/l. During operation time of 7 years 80% of the nitrate was removed from the wastewater. At the Borden site, nitrate removal efficacy of the sawdust barrier was 74% since installation in 1992. The average loading rate was approximately 200 l/day containing 28 mg/l N (nitrate and ammonium). The residence time is about 15 days. A vertical denitrification barrier intercepting a horizontally migrating septic system plume was installed at the Long Point site in Canada. The barrier consisting of 20% of coarse hardwood sawdust shows a removal efficacy of 91% of nitrate during monitoring period (since 1992). The influent nitrate-N concentration is about 34 mg/l. The average flow rate is 15 l/day resulting in an average residence time of 13 days. At the North Campus Site in Waterloo, Canada, a barrier was installed as a containerised subsurface reactor treating farm field drainage water. The reactive material consists of 100% coarse wood mulch. The hydraulic loading range from 800 to 2000 l/day resulting in residence times of 3 - 7 h. The input nitrate concentration is 4.8 mg/l, the removal efficacy is 58% since 1993. The denitrification barrier installed at the Bardowie Farm, Cambridge, New Zealand, worked successfully for a period of 2.5 years (Schipper and Vojvodic-Vukovic, 2000). The continuous barrier consists of 50% sawdust mixed with the excavated soil. Input nitrate concentrations of 5 - 1 6 mg/l were measured. The nitrate removal from contaminated groundwater ranges from 0.8 to 12.8 ng/cm3hr, depending on groundwater table (seasonal variations). Generally, the nitrate reaction rates are temperature-dependent ranging from 5 mg/l per day (2-5~ to 15-30mg/l per day (IO-2O~ with no indication of deteriorating performance during barrier operation time. The limiting factor on long-term performance of denitrification barriers is obviously the availability of carbon. Mass balance calculations show that the carbon consumption by heterotrophic denitrification ranges from 0.6 kg at the Killarney to 4.5 kg at the Borden site, respectively, which represents only 2 - 3 % of the initial carbon mass at each site during operation time. Carbon consumption by other reactions like reduction of dissolved oxygen (DO), DOC leaching and sulphate reduction have little influence on the long-term performance of these systems. The results show that organic carbon barriers can be designed to provide 10 years or more of nitrate remediation without replenishment of the reactive material. GAC is mainly used in Europe in sorption barriers for the treatment of organic contaminants. At the Bitterfeld site in Germany, in situ reaction wells filled with GAC have been installed to remediate organic solvents from contaminated groundwater (Kraft and Grathwohl, 2003). Under in situ conditions, the sorption capacity of the activated carbon is strongly influenced by the hydrochemistry of the groundwater. Surface coatings may decrease the sorption capacity and the sorption kinetics resulting in earlier breakthrough of the contaminants. Preliminary column tests with activated carbon showed that after contaminant treatment (1200 pore volumes) the activated carbon was coated with


19

various minerals (Kraft et al., 2000). Three different types of GAC were investigated during an operation time of 2 years (Kraft and Grathwohl, 2003). The groundwater is predominantly contaminated with monochlorobenzene (MCB) with concentrations up to 25 mg/l. The treatment of MCB by each GAC filtration system was successful and efficient during operation time. The predicted MCB loading of the GAC by modelling showed a very good agreement with the measured contaminant breakthrough in the in situ reaction wells, ranging from 9900 to 16,400 pore volumes depending on the type of GAC. These results show that the expected sorption capacity was indeed achieved by the GAC, proving a sufficient long-term performance of GAC barriers. After a certain equilibration period (100 pore volumes) the groundwater chemistry up- and down-gradient the reactive wells was stable. During the monitoring period in none of the activated carbon reactors plugging, chemo- or biofouling was observed, as documented by an unmodified groundwater chemistry (pH values ranging from 6.6 to 6.9, anaerobic conditions). The particle loss of the filter systems of 0.02% had no measurable influence on the hydraulic properties of the GAC. Compared to fresh GAC the specific surface area of treated GAC is decreased by up to 6-26.5% due to the groundwater constituents adsorbing on the GAC surface. The use of a sequential system of TCE degradation by Fe~ followed by the adsorption of monochlorobenzene (MCB) on GAC was investigated in column tests by K6ber et al. (2002a). The results showed that the lifetime of the activated carbon could be increased by a factor of 4 in combination with Fe ~ compared to pure GAC. The potential decrease in GAC sorption capacity due to surface mineral precipitation can be minimised or prevented by separation of the reduction zone and the adsorption zone.

E. Outlook

Not all the PRBs installed in recent years are true success stories, a fact that can be deducted from the lack of publicly available information on some of the pilot-scale and full-scale installations. Most problems appear to be related to the system hydraulics, with the geochemistry of the contaminant attenuation process and the reactive material actually functioning as predicted. The hydraulic functioning of a PRB relies to great extent on the understanding of the local and regional aquifer systems, the planning (hydraulic modelling) of the PRB system to be installed, and the quality of the construction work (Parbs et al., 2003). Therefore, future improvement of the PRB technology needs to address these issues. Other fields of advancing the PRB technology are the study of new and innovative materials targeting specific contaminants, and the combination of PRBs with other remediation technologies such as bioremediation (Werner, 1998; Scherer et al., 2000; Tiehm et al., 2000), monitored natural attenuation (Carey et al., 2002) or electrokinetics (Ho et al., 1995; Chew and Zhang, 1998; Czurda and Haus, 2002).

References

Admassu, W., Breese, T., 1999. Feasibility of using natural fishbone apatite as a substitute for hydroxyapatite in remediating aqueous heavy metals. J. Hazard. Mater. B69, 187-196.

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Long-term Performance of Permeable Reactive Barriers K.E. Roehl, T. Meggyes,F.-G. Simon, D.I. Stewart, editors 9 2005 Elsevier B.V. All rights reserved.

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Chapter 2 Construction methods of permeable reactive barriers Tam~s Meggyes

A. Introduction

Construction methods for permeable reactive barriers (PRBs) have been developed using experience gathered with cut-off wall construction techniques due to many similar features between the two technologies (Meggyes and Simon, 2000; Simon et al., 2002a). First, cut-off wall construction techniques were applied to PRBs without almost any alteration. Single- and two-phase diaphragm walls, bored-pile walls, jet grouting, thin walls, sheet-pile walls, driven cut-off walls, injection and frozen walls are the most common cut-off wall alternatives. To date, in addition to using cut-off wall construction methods, an increasing number of innovative techniques are being used to construct PRBs such as drilling methods, deep-soil mixing, high-pressure jet technology, injected systems, column and well arrays, deep aquifer remediation tools (DART), hydraulic fracturing and biobarriers. The main configurations of PRBs are: 9 9 9 9

continuous reactive barriers, funnel-and-gate systems, arrays of wells, injected systems.

B. Construction of cut-off walls

The most commonly used cut-off wall construction methods apply one of the following alternatives (Brauns, 1994; Meggyes and Pye, 1995; Meggyes and Simon, 2000; Simon et al., 2002b): 9 Trench excavation using supporting fluids capable of solidifying and forming a diaphragm wall (single-phase diaphragm wall) or which - after excavation of the trench - are displaced by another material which, in turn, is capable of solidifying (twophase diaphragm wall). 9 Forming a thin slot by driving a beam into the ground, then consecutively retracting the beam and filling the space with a thick slurry (e.g. thin walls). 9 Driving strong elements into the ground (e.g. steel sheet-piles).

T. Meggyes

28

9 Constructing interlocking boreholes and backfilling them with concrete to form a wall of interlocking columns. 9 Injecting or placing reactive materials into the ground in a discontinuous fashion. Cut-off walls can be used to prevent pollution migration in any of the following cases: 9 where there is contact between the contaminant and the groundwater and/or flow through a contaminated body; 9 where there are mobile liquid pollutants above the groundwater level; 9 to prevent migration of gaseous pollutants in the unsaturated zone; 9 as a hydraulic control measure; 9 to contain a reaction space for in situ decontamination. Since the selection of the most suitable cut-off wall construction technology depends on a large number of local conditions and other aspects, not all construction methods will be necessarily used in reactive barriers. Nevertheless, the aim here is to briefly describe as wide a range of cut-off wall construction methods as possible, so as not to exclude any from future consideration. The most commonly used cut-off wall systems are illustrated in Table 2. l.

1. Cut-off wall types 1.1. Single-phase diaphragm wall When constructing a single-phase diaphragm wall, 0 . 4 - l m thick panels are excavated from the soil using grab buckets, clamshells or vertical trench cutters (Arz, 1988). A selfhardening slurry is pumped-in to stabilise the trench walls and form the final wall. Cut-off walls are constructed using the "pilgrim' s pace" method: i.e. primary panels 1,3,5, etc. are excavated and filled with slurry first. After a period of 3 6 - 4 8 h, as soon as the slurry in the primary panels has hardened to a cuttable state, work begins on the secondary panels 2,4,6, etc. The cutter excavates the secondary panels with an overlap of 0.3-0.6 m cutting into the primary ones. Since the hardening process in the primary panels has not yet finished, an intimate contact is achieved between the primary and secondary panels. This feature has a positive advantage over the two-phase method where gaps are created by removing the stop-end tubes. Also, using the single-phase method it is not necessary to dispose and replace the slurry, thus no typical fault lines are created. Imperfections can considerably impair barrier performance: a 1 m 2 hole lets as much groundwater escape as a 100,000 m 2 high-quality cut-off wall does (Dtillmann, 1999). Various types of misalignment, such as tilting, turning or twisting of the panels may result in insufficient overlapping and, as a consequence, in an increase in the system's permeability (Stahlmann and Scholz, 2004).

1.2. Two-phase diaphragm wall Two-phase diaphragm walls are constructed in two steps. In the first phase, soil is excavated while a bentonite suspension stabilises the trench walls. In the second phase, the bentonite suspension is replaced by the cut-off slurry using tremie pipes. The individual panels are confined by stop-end tubes. Problems and deficiencies may arise when long

Table 2.1. Cut-off wall systems (Jessberger, 1992). ~

Principle Excavation of soil and placement of sealing material

cut-off wall system

Soil

Single-phase dlaphrdgm Wall

Limited suitability for peat and humic acids Limited suitability for peat and humic acids Limited suitability for peat and humic acids No limitation if casing is applied

Two-phase diaphragm wall Composite diaphragm wall Interlocking bored-pile diaphragm wall Displacement of soil and installation of sealing material

Thin wall

Sheet-pile wall

Suitable for pile dnving

Driven cut-of wall Reducing permeability of soil in place

Injection wall

Injectable

Jet grouting, HPI

Very fine-grained

Frozen wall

Material

Depth (m)

Thickness (m)

Permeability

Bentonite-cement mix with or without filler Bentonite slurry, soilcrete

Approximately 35

0.4- 1.5

51x10

>50

0.4- 1.5

0.6

Approximately 20

0.6-0.8

< 5 x 10- '(' for cut-off slurry < 5 x lo-'"

Bentonite- cement mix with fillers

18-23

0.05-0.20

I1

Steel

15-20

0.01 -0.02

-

Soilcrete, concrete

15-20

>0.4

5 1 x 10-9

Cement, achy -cement mix, silicate gels Bentonite-cement amix with and without filler Liquid nitrogen

> 100

adjustable

51x10

> 100

>0.8

5 1x

> 100

>0.8-1

.o

Experience

tdS) I('

10-10

x lop9

10-l0

Landfill 28 years

Landfill 20 years

Landfill 20 years Landfill 14 years

Landfill 18 years

Contaminated land 18 years Contaminated land 19 years Hydraulic and foundation engineering Hydraulic and foundation engineering Civil and mining engineering

30

T. Meggyes

stop-end tubes are retracted and the bentonite suspension is replaced by the cut-off slurry. In order to ensure an efficient replacement it is essential that the cut-off slurry has a density exceeding that of the bentonite slurry by 500 kg/m 3. If the bentonite suspension is not completely replaced, faults remain near the joints in an otherwise watertight wall.

1.3. Composite cut-off wall In both, single- and two-phase systems, additional elements can be inserted into the cut-off wall to improve strength and/or water-tightness. Sheet piles and glass walls/tiles are a few examples, with geomembranes being used most frequently. These elements are placed into the fresh bentonite cement mix immediately after placement. Because of their lightweight, a special construction apparatus is required to insert geomembranes into the cut-off wall and special locks provide watertight joints between the plastic geomembrane sheets. The locks can be welded and checked after the cut-off slurry has hardened. Glass walls in composite cut-offs provide a high-degree of water-tightness, however, their length is limited and the problem of forming a perfect watertight seal at the joints remains to be solved. Sheet pile walls can also be combined with bored piles around joints (Jessberger and Geil, 1992; Diillmann et al., 1993; Meseck, 1987; Stroh and Sasse, 1987; Ghezzi et al., 1999; Fischer, 2002). Multiple-layer barrier systems have also been suggested (Cavalli, 1992): they consist of three walls with increasing chemical resistance and lower permeability. The components are: an outer 3-mm thick bentonite filter cake, a O.3-O.6-m thick soil-bentonite, cementbentonite or plastic-concrete middle layer and an inner 2.5-mm thick HDPE geomembrane with a permeability of 10 -1~ m/s. Construction starts with excavating the trench under a bentonite and/or cement slurry, the geomembrane is then installed mounted on a detachable and removable frame using weights or a pile driver. Once the HDPE is in place, the trench can be backfilled with an option to install a monitoring system.

1.4. Cut-off wall chamber systems A cut-off wall chamber system consists of two parallel cut-off walls a few metres apart, connected to each other by cross walls about every 50 m and usually combined with the removal of water by pumping from within the chambers to increase safety. The Rautenweg landfill in Vienna is an example where the geological sealing layer lies at a depth of 7 0 100 m (Arz and Weber, 1987; Brandl, 1989). The parallel cut-off walls are 8 m apart and the connecting cross walls are arranged at every 5 0 - 7 0 m. The cut-off wall chambers formed in this way have a surface area of 400-600 m 2, and in each of the chambers a well was constructed and a pump installed to maintain a groundwater level 0.5 m below the natural level outside the landfill. The groundwater beneath the landfill is kept at 0.2 m below the level inside the chambers. This enables checking the water-tightness of each chamber after completion by pumping tests. It is sufficient to sink the cut-off walls to a layer of reduced permeability (e.g. k f - - 5 x lO-6m/s) which enables an optimal maintenance of the groundwater level by continuous pumping. In another large-scale application, a 3.7-km long cut-off wall chamber system was constructed to contain the landfill in Vorketzin near Berlin where large amounts of waste from the former West Berlin had been deposited (Kellner and Scheibel, 2004). The two

Construction methods of permeable reactive barriers

31

0.6-m thick parallel single-phase cut-off walls run 3.6 m apart and are connected with cross-walls at 5 0 - 1 8 0 m intervals. Water is pumped from within the chambers and transported to a leachate treatment plant. The top of the inner cut-off wall is lower than the external one and allows for the leachate to overflow into the chambers should the water level rise within the landfill. The "pilgrim's pace" method was used and the secondary panels were constructed 3 days after the primary ones when strength in the primary panels provided a cuttable state. The construction sequence of the three walls was as follows: (l) cross walls, (2) inner cut-off wall, (3) external cut-off wall. Solids content of the cut-off wall material was 450 kg/m 3 and a part of the material contained special additives to counteract delay-effects of ash within some of the soil. A total of 80,000 m 2 of cut-off wall was constructed consisting of 2250 panels. Quality assurance included more than 16,000 tests of various types. Uniaxial compressive strength was 1042 and 1732 MN/m 2, and permeability 1.4 x 10 -10 and 1.7 x 10 -10 m]s.

1.5. Thin walls First sheet piles, then heavier steel beams are vibrated into the ground and a claycement-water mix is injected into the void as the beams are retracted. The panels are cut into the adjacent ones, so that there is an overlap and water-tightness is ensured (Arz, 1988). The toe of the beam is also provided with a blade to guide the beam as it cuts into the previously constructed panel to provide the desired overlap. Feasible depths for thin walls are between 15 and 20 m. The thickness depends on the dimensions of the former at the toe which is usually 6 0 - 8 0 mm thick. Greater thickness increases the driving resistance and the retraction force, thus thicker formers are seldom used. Wall thickness is also influenced by the soil layers and injection pressure. In coarse layers (sand, gravel) the vibration has the additional effect of compacting the surrounding medium thus achieving a reduction in permeability. The clay-cement mix usually penetrates into the pores of the surrounding soil, especially when loose sediments dominate; thus, the final wall thickness achieved is greater than the nominal thickness. Driven thin walls are very economical cutoff wall systems. This method enables 400-1000 m 2 of cut-off to be installed per shift. To achieve an impermeable wall, very accurate pile guiding and control by precise instrumentation is required.

1.6. Sheet-pile walls Sheet-pile walls are usually constructed using steel piles driven into the ground, though precast concrete, aluminium or wood piles can also be used. Steel sheet-pile walls are easy to construct, can carry heavy loads (which may be of benefit if they are used to form box-type gates in funnel-and-gate reactive barriers), their construction time is short, little room is required for their construction, and there is no need for contaminated soil to be disposed of (Roth, 1988; Weber et al., 1990; Jessberger and Geil, 1992; Rodatz, 1994; Berndt, 2002). Other advantages are that containment is achieved immediately after construction, and as the steel sheets are manufactured in factories quality control is good. Joint-sealing to prevent leakage through interlocking sheet piles was a problem in the past because of leakage, but innovative solutions introducing sophisticated labyrinth joints and contaminant-resistant

32

T. Meggyes

sealing pastes or plastic sealants (Hoesch) are now capable of providing a high degree of water-tightness. The locks connecting the sheet-piles can be checked both electrically and by measuring the pressing force (Schultze and Mul3otter, 2001).

1.7. Bored-pile cut-off walls and jet grouting Bored-pile cut-off walls are constructed with secant piles. After sinking the primary piles 1,3,5, etc., the secondary ones 2,4,6, etc. are constructed in such a way that they are cut into the primary piles, thus establishing an intimate contact between them. Cut-off walls can also be constructed of soilcrete columns, also called jet grouting (T6th, 1989; Kutzner, 1991; Jessberger, 1992): a rotary drilling technique is used in this technology and a high-density mud serves both as a cutting medium and filling fluid. Contaminated underground bodies can be treated using a novel two-phase jet grout system (Dwyer, 1998) so that contaminated spoil remains in the subsurface. First air or water is injected at high pressure to form a hollow underground storage cavern above the contaminated body and clean soil from the cavern is removed to the surface. In the second step, the contaminated body is stabilised while contaminated drill cuttings are deposited in the hollow cavern formed in the first step.

1.8. Injection walls This method reduces soil permeability by injecting a solidifying liquid through boreholes into the pores and fissures of the ground. The most frequently used materials injected are cement suspensions, artificial resins or water glass-based materials. The distance between the injection holes is determined by the rock permeability, the viscosity of the injected material and the highest permissible injection pressure (Kutzner, 1991; Jessberger, 1992; Schulze, 1992; German Geotechnical Society, 1993).

1.9. Frozen walls Frozen walls were considered a temporary solution for a long time, but their technology has made considerable progress and they are now increasingly being used as permanent solutions. A closed, watertight body can be produced by inserting pipes into the ground and circulating a refrigerant or liquid nitrogen in them. Latest developments suggest that permeability of a frozen soil can be as low as 10-12 m/s and diffusivity around 10 - 9 cmZ/s (Dash et al., 1997; Mageau, 1998), especially when advanced methods with thermistor or electro potential monitoring are used. Installation of frozen walls requires little or no soil excavation and the walls can be removed by stopping the cooling (although they can last months without power, improving their energy-efficiency) and there are no wastes for disposal. The feasible depths are around a few hundred metres and both clay and sand/ gravel type soils may be targeted.

2. Cut-off wall materials The water-tightness of a cut-off wall is basically determined by the properties of the cut-off slurry used. Mineral cut-off slurries usually consist of bentonite, cement, fillers and water,


33

although in special cases chemical additives are also used. The composition and the properties of mineral cut-off slurry have to be determined by suitability tests for each construction task, and construction should usually be controlled by appropriate quality assurance measures (Meseck, 1987; Jessberger and Geil, 1991). Since it is common practice to use local and non-standard materials, the results of suitability tests are usually difficult to apply to other projects. Cut-off slurries can be distinguished according to the type of cut-off wall: 9 thin wall slurries; single-phase slurries; 9 two-phase slurries.

9

Thin wall slurries must have a relatively high solids content to ensure a high density. To prevent the closure of the void while the driving pile is being retracted, a density of 1600 kg/m 3 is required and the cut-off slurry must remain pumpable even at this highsolids content. A well-proven mixture for thin wall slurries is (Arz, 1988) 9 9 9 9

25 kg bentonite 175 kg Portland cement 800 kg rock flour 6401 water

Cut-off walls of this composition have a permeability of approximately 10 -8 m/s after setting. Hardened cut-off slurries for single-phase walls must be machinable (i.e. their yield stress must not exceed 80 N/m 2) otherwise satisfactory interlocking of the panels cannot be achieved. On the other hand, they must provide a high degree of water-tightness, which, in turn, requires a high-solids content (bentonite and cement). These somewhat contradictory requirements can be reasonably satisfied (kf----5 • 10 -9 m]s) using the following mixture: 42 kg Na-bentonite 9 200 kg kiln cement 9 917 1 water

9

Calcium bentonites are also used to produce cut-off wall slurries with high solids content. A well-proven mixture is 1 6 5 kg Ca-bentonite 9 144 kg kiln cement 9 826 1 water

9

Ca-bentonites exhibit lower swelling and provide more stable suspensions than Nabentonites. Disadvantages are a high water loss and a stiffening of the mix which can be compensated for by adding modifying chemicals; however, these must not react with the groundwater. Much higher solids content can be used for two-phase cut-off wall slurries (Hitze, 1987). These cut-off slurries contain clay, bentonite and cement, with added fillers (e.g. sand and gravel) and can have a density of up to 2000 kg/m 3 (Seitz, 1987). Two-phase cutoff slurries provide permeabilities of kf-- 5 X 10-11 ITI/S.

34

T. Meggyes

The mechanical properties of hardened cut-off materials depend on the hydraulic binder, solids content and time. In practice, the uniaxial compressive strength is measured shortly after the cut-off material has hardened (using standard tests for measuring the compressive strength of soil) and after 28 days (using standard concrete technology methods). The compressive failure strain of 28-day samples tends to lie between 1.0 and 2.0%, breaking strength is 500-1000 kN/m 2. Fourteen-day-old samples with low cement content show plastic behaviour beyond the compressive failure strain. Young's modulus lies in the range 50-lOO MN/m 2 and shear strength in the range 100-500 kN/m 2. Strength measurement is widely used to monitor cut-off wall construction, although strength is only really important when the wall is used as a load-bearing structure, and ductility is the main property required from cut-off wall materials. The justification for monitoring strength gain is that measured trends can be compared with expected trends, to demonstrate that the correct mixture at the desired density has been achieved at a particular location. The problem is that too many engineers may start to consider that high strength is good whereas a slow rate of strength gain and high strains to failure are the desirable properties. Chemical resistance over the planned design life is an important feature of hardened cut-off wall slurries (Dietrich et al., 2004). Apart from the results of suitability tests, knowledge on the concentration-dependent impact potential of the most frequently occurring contaminants is of utmost importance in order to be able to optimise cut-off wall materials. In this context information borrowed from concrete technology is often useful. A magnesium concentration of 250 mg/l can exert a strong chemical impact on single-phase cut-off wall materials. The damage mechanism is mainly triggered by an extensive precipitation of magnesium hydroxide, which reduces the pH value in the pore water. Dietrich et al. (2004) found that the rate at which the thickness of the softened layer (shown by uniaxial strength values) increases with the magnesium concentration. The penetration depth of a Vicat needle can fairly accurately register the interface between solid and softened-up layers. Sulphate, however, is much less damaging: concentrations as high as 15,000 mg/l failed to cause serious damage to single-phase cutoff wall materials, which can be explained by their low tricalcium aluminate content and low permeability. An increase in solids content proved beneficial for the chemical stability against carbonic acid, while addition of limestone flour has not changed stability (Dietrich et al., 2004). The sensitivity of cement to some chemicals has led to the development of cement free cut-off slurries for two-phase walls (Seitz, 1987), these contain powdered clay, water glass, sand, gravel and a silicone reagent. They have densities as high as 2300 kg/m 3 and permeabilities as low as k f - - 1 0 -11 m]s. Their stiff-plastic behaviour is highly advantageous, especially if subsidence can be expected (Hitze, 1987; Jessberger and Geil, 1991).

C. Construction of reactive barriers

The basic performance requirements for a reactive zone within a PRB are (Beitinger and Btitow, 1997; Smyth et al., 1997; Beitinger, 1998):

Construction methods of permeable reactive barriers 9 9 9 9

35

replaceability of the reactive materials; higher permeability than the surrounding reservoir; stability against fines washed into barriers from the surrounding soil; long life-span.

The selection of the construction technique mainly depends on site characteristics (Gavaskar, 1999), e.g.: 9 depth (the most important factor): increasing depth requires more specialised equipment, longer construction times and is accompanied with higher costs; 9 geotechnical considerations: soil/rock strength and presence of obstacles; 9 soil excavation: handling and disposal of (contaminated) soil; 9 health and safety during construction (entry of personnel into excavations). In cases with depths of less than 8 m a trench is usually excavated and simultaneously filled with the reactive material (Puls, 200 l). In more complex cases (Fig. 2. l) the reactive barrier consists of the core part filled with reactive material and filter gravel preventing fines from the soil from entering the reactive zone. The top of the reactive barrier is covered with low-permeability materials (clay), which excludes any contact with oxygen from the air. The overall permeability of the barrier should usually be 50-200 times greater than that of the surrounding soil to encourage groundwater to flow through the barrier instead of by-passing it.

1. Design considerations The design of PRBs, as with any other technology, should meet the requirements of the best available technique (BAT). Thus, design must include adequate data gathering, and an evaluation process in which all-appropriate technical, ecological and economic criteria are carefully and objectively considered. In most cases, it may be advantageous to develop

Figure 2.1. Structure of a reactive barrier (Beitinger and Btitow, 1997).

36

T. Meggyes

a reliable conceptual site model and to perform pumping and treatability tests. Operational issues such as long-term performance, efficiency, costs and monitoring have to be analysed. Site-specific assessments for the chosen technology are advantageous because they encourage better acceptance by relevant authorities and affected neighbours. It should not be assumed in advance that a particular innovative technology is the best solution; it should always be shown on a site-specific basis that the selected remediation solution represents the Best Available Technique Not Entailing Excessive Cost (BATNEEC). So-called BAT reference documents (BREFs) can be downloaded from the BREF site of the IPPC Bureau (http://eippcb.jrc.es/pages/FActivities.htm) (Beitinger, 2002). On most contaminated sites, data gathering is focused on the identification of contaminants and the delineation of plumes and/or free phase volumes in the unsaturated and saturated zones below the spills. To determine potential remediation alternatives, additional data must be gathered and evaluated in respect of the hydrogeology, hydrochemistry and migration of dissolved and undissolved chemicals (Table 2.2). Indeed, this process should include those data which identify potential problems with treatment technologies, such as precipitation of iron, manganese, calcium and magnesium and bioclogging. The general objective of data gathering is to fully understand all criteria with potential to either support or exclude particular treatment alternatives. Further, it is suggested that a so-called conceptual site model should be developed for any site where treatment is proposed, as this will facilitate the understanding of contaminant distribution, migration, adsorption, degradation, convection, diffusion and all other transport and retention mechanisms, as well as chemical reactions and physical behaviour. Borehole log evaluation, groundwater monitoring, extensive pump tests and geophysical investigations are the common methods of choice for the gathering of reliable hydrogeological data. Heterogeneous underground structures in the unsaturated and the saturated zones must be carefully surveyed. Geochemical and hydrochemical data are of major importance to evaluate migration and degradation processes in situ as well as treatment technologies in situ and above ground (the latter assuming pumping). Groundwater specimens should be collected by pumping, and the water should only be sampled when pH-value, electrical conductivity, redox potential, oxygen content and temperature show constant values (Beitinger, 2002). Monitoring the spreading of contaminants over time will help elucidate emission rates, spill migration velocities and contaminant retardation, as well as any degradation processes that may be occurring. The conceptual design report shall include the following information (Beitinger, 2002): 9 The amount and type of any emissions from the remediation scheme and details of any emissions control measures. 9 The volume and water quality of any discharge or re-infiltration of treated groundwater. 9 The power consumption (electricity, fuels, etc.). 9 A description of any waste streams generated and details of their disposal. 9 Quantification of any material inputs such as GAC, lime, etc. 9 The remediation target levels. 9 The anticipated overall efficiency. 9 The anticipated maintenance requirements (manpower, parts). 9 The monitoring requirements.


37

Table 2.2. Data for design purposes (Beitinger 2002). Remarks Hydrogeological data Hydrogeology Depth to groundwater table Aquifer thickness Groundwater flow direction Hydraulic permeability Groundwater gradient Transmissivity Confined aquifer? Surface water bodies Weather conditions Surface conditions Geochemical and hydrochemical data pH-value of soil and water Electrical conductivity, TDS (salinity of water) Redox potential Oxygen content Temperature Iron Manganese Calcium Magnesium Carbon dioxide Sulphate, sulphide Nitrogen, Total Kjeldahl Nitrogen (TKN) Other chemical compounds BOD, COD Contaminant distribution Identification of contaminants Delineation of plume Plume activity Free phase spreading Residual saturation (Sr) DOC (Dissolved Organic Content) TOC (Total Organic Content) Potential receptors Age of pollutants Migration with time Spill location/contaminant sources

General description of geology, aquifers/aquitards, anomalies (m) (m) (k f in m/s) (J) (T) Pressure Description, distance Precipitation rates, wind factor Surface covers, plants, asphalt

Hardness Hardness Precipitation of Ca, Mg Potential inhibitors Nitrate, nitrite, ammonium "Background levels", metals Biological and chemical oxygen demand Types of pollutants Area, depth, concentration in soil and groundwater Increasing/decreasing, time factor LNAPL/DNAPL Unsaturated zone Concentration of dissolved organic matter in water (incl. suspended particulate matter in water) Identification, distance, sensitivity Ageing/degradation processes Points of emission

T. Meggyes

38

Table 2.2. (continued) Remarks

Contaminant properties and transport characteristics Density LNAPL/DNAPL Liquid viscosity Interfacial tension with water Solubility Vapour pressure Henry's law constant Partitioning coefficient (Kd) Organic content in the soil (foc) Octanol/water partition coefficient (Kow) Organic carbon partitioning coefficient (Koc) Ion exchange capacity Clay fractions Biodegradability Grain size distribution Bulk density of aquifer material Air permeability in soil Soil vapour Soil porosity Water content Soil heterogeneity

9 A detailed cost estimate (including the capital costs, construction costs, operating costs and decommissioning costs). 9 A detailed health and safety evaluation of the project. Table 2.3 shows the common cost parameters that should be considered for cost estimation purposes, and a cost example is given in Table 2.4.

2. PRB construction technologies 2.1. Cut-off wall technology for PRBs The methods applied in cut-off wall excavation as described in Section B can be used for reactive barriers. Reactive materials can be placed into the trench using common earthmoving machines. Indeed, the funnel elements in funnel-and-gate systems are conventional cut-off walls. Steel sheet-piles are increasingly being used as parts of PRBs due to their advantageous features (Schultze and MuBotter, 2001; Morrison et al., 2002). Gravelding (1998) reports on a 400 m long funnel-and-gate barrier that has been constructed using steel sheet-piles (vibrated into the ground) as the impermeable "funnel" walls. In the Karlsruhe East gasworks site remediation project a 240 m long and 19 m deep steel sheet-pile wall was built to form the funnel element for a funnel-and-gate system (Schultze and Mul3otter, 2001). The sheet piles were pressed into the ground using a silent piler to avoid damage to nearby historic buildings.


39

Table 2.3. Cost parameters (Beitinger 2002). Capital expenditure

Purchase cost (including taxes) Mobilization/installation Start-up costs Interest rate Operational and maintenance costs

Personnel costs Energy costs (electricity, fuels) Consumable materials (GAC, lime, etc.) Maintenance Monitoring (sampling, chemical analysis, reporting) Discharge costs Residuals, waste disposal costs Years of operation Fees and taxes

2.2. Gate structures

Gates are the crucial elements of funnel-and-gate systems. A frequently applied gate structure uses sheet-pile boxes in which either the sheet piles are perforated, or suitably arranged gaps provide access for the groundwater. Sheet piles can also be used to separate soil, gravel, reactive material, etc. during placement and then retracted without disturbing the arrangement of the different materials. In the example of the 400 m long funnel-and-gate barrier reported by Gravelding (1998), four 12 m wide cells containing iron served as gates and were constructed as sheet pile boxes. The bottoms of these boxes were lined with bentonite and a geotextile, upon which iron filings surrounded with pea gravel were placed (the pea gravel to provide assisted groundwater entry and egress access for the groundwater). The gates were capped by geotextile and bentonite and native soil was backfilled to grade. The residence time (determined from the influent chemistry, reaction rates and treatment goals) combined with the groundwater flow velocity meant that 0.6-1.8 m cell thickness (containing 580 tons of iron) were required. The barrier installation amounted to US$200-250 per m 2, gate installation costs were US$850-lOOO per m 2. In the Karlsruhe East gasworks site remediation project a funnel-and-gate system was built to clean up polycyclic aromatic hydrocarbon (PAH) and benzene contamination (Schultze and Mu6otter, 2001). The gates were constructed as large-diameter (2.5 m) wells and filled with activated carbon (Fig. 2.2). 2.3. Reactive thin walls

These apply the thin wall technique developed for cut-of walls (described in Section B. 1.5) and unite the advantages of slurry trenching and sheet-piling (Jansen and Grooterhorst, 1999). A hollow steel beam is vibrated into the ground to create the space for the wall, and

40

T. Meggyes Table 2.4. Cost example (Beitinger 2002).

Dimensioning and Cost Criteria

PRB

Aquifer Permeability Hydraulic gradient Pore volume Groundwater capacity Installations

10m 10 -4 m/s 0.001 0.2 20 m3/h 3 gates (10 • 15 m), funnel length: 270 m

Contamination CHC GAC loading GAC volume Loading time

1 mg/1 1% CHC/weight 150 m 3 --~3.5 years

Investment Depreciation time Depreciation costs Interest rate Interest costs

1 million C 20 years 50,000 C/year 5% 50,000 C/year

Total capital costs

100,000 C/year

Electricity (l 0 W/h) Operation (manpower) GAC Repair- maintenance Discharge of groundwater (0.50 C/m3) Monitoring

40,000 C/year l 0,000 C/year

Total operational costs

60,000 C/year

Total costs Specific costs per treated volume (175,200 m3/year)

160 000 C/year 0.91 C/m3

m

10,000 C/year

a reactive material is placed into the void as the beam is retracted. A number of "panels" constructed in this way next to each other form a continuous reactive wall. The main advantages of this method are low demand on space, no soil or water extraction, minimum impact on groundwater flow, possible depths of around 25 m and the reactive material can be recovered using the same hollow steel beams. It must be noted, however, that small thicknesses used allow only a short residence time making reactive thin walls suitable only for cases when fast contaminant removal can be achieved.

2.4. Drilling and deep soil mixing

Contiguous circular columns containing reactive material can be installed using either drilling or deep soil mixing (Day et al., 1999). Longer barriers are formed by inserting columns at closely spaced centres. Caisson drills can be used for the drilling, with column


41

Figure 2.2. Cross-section of a gate (Schultze and MuBotter, 2001).

diameters of 0.5-2.5 m. Usually, a large circular casing is lowered into the ground, the soil is removed by augers and the hole filled with a reactive material. Instead of casing, biodegradable polymer slurries (Hubble et al., 1997) or shear-thinning fluids (Cantrell et al., 1997) may be used as supporting fluids with significant cost benefits. Alternatively, deep soil mixing blends the soil with slurry in situ without excavating the soil (Gavaskar et al., 2000). A caisson is lowered into the ground, a set of multiple augers penetrates the ground and the reactive material is injected through the hollow kelly bar of the mixing tools. The reactive material is injected in slurry form, which is then mixed with the soil by the augers. The soil remaining in situ “dilutes” the reactive material to 40-60%. 2.5. Jet technology

The proposed technique involves using high-pressure jets and jet pumps (Debreczeni and Meggyes, 1999). It is envisaged that a jet cutting head, joined to the drilling pipe by a hinged connector and a flexible hose, is used for trench excavation (see Fig. 2.3). This allows the drilling pipe to be swung from side to side within the trench by means of a hydraulic mechanism so as to achieve the appropriate length of excavation. Soil excavation in low-strength strata can be performed by slurry jets and in high-strength strata by a milling head, which can be attached to the drilling pipe and operated by a hydraulic motor. The cuttings can be transported to the surface through the drilling pipe by means of a jet pump. Where slurry jets are used in the trench excavation, both the jet cutting head and the jet pump can be operated by the same slurry pump. Depending on construction process and final barrier requirements, the slurry can either remain in the trench after excavation is

42

T. Meggyes Hoist

Arch or drum screen Drillings

I[

Slurry pump

' ,Jl

",l

, illli

! - -

I

,, _--5

'

I

,

--

\ Jet cutting head

Stop-end tube

I

Figure 2.3. High-pressure jet equipment proposed for cut-off wall or reactive barrier construction.

complete and form the final barrier (single-phase technique) or will be replaced by the final barrier material (two-phase technique). After being pumped to the surface the drillings must be separated out from the slurry, which can then be re-circulated. Solid particles greater than 1 mm can be removed by arch or drum screens, for finer materials a hydrocyclone is required. The cutting head and the drilling pipe can be carried by a hoist or drilling mast or supported by stop-end tubes or sheet piles which can be removed and re-used once the panels have been completed. Depending on local conditions, especially where mechanical cutting is used, air lift can be considered instead of jet pumping. A combined jet method has been suggested where iron particles are used both as a reactive material and abrasive medium in trench excavation. The granulated iron particles can be entrained in high-velocity water jets and erode the soil or rock. This technique enables trench construction and placement of the reactive material simultaneously. The barrier thus constructed is a hybrid of a reactive wall and an injected system. 2.6. Well-based systems or DART

Arrays of wells also known as the DART represent an alternative to continuous barriers (Wilson and Mackay, 1997; Golder Associates Ltd., 1998; Freethey et al., 2002).


43

This may be helpful at sites where installation of permeable barriers may be impractical for technical or financial reasons. An array of wells can form the gate within a funneland-gate system or serve as a set of in situ reactors or can release substances that encourage biological or other processes. An array of wells is essentially a noncontinuous reactive barrier with the advantage that drilling technique can be used to construct them, allowing lower installation costs and greater depths. The wells filled with reactive materials have lower hydraulic resistance than the surrounding soil, thus the groundwater flow converges towards them. The amount of reactive material required for the well is much less than for a continuous reactive barrier. An important consideration is whether such a barrier can achieve sufficient residence time for effective treatment, which will depend primarily on the groundwater flow-rate. The whole contaminant plume must flow through the array of wells and thus the optimum well spacing is of great importance. As a rule of thumb, well spacing should be about twice the well diameter. If the required residence time is high, then more than one array of wells may be required. When a treatment compound is released from the barrier, the wells may be less closely spaced. The DART is essentially a further development of well-based systems. Its application further emphasises the following advantageous features: 9 9 9 9 9

Materials, equipment and contractors are easily available. Well installation is straightforward. Greater depths (> 50 m) are feasible. Only a fraction of the reactive material is needed in comparison to continuous walls. Replacement of spent reactive material is fairly easy and inexpensive.

A DART well is composed of a rigid PVC casing with high-capacity flow channels that contain the reactive material and flexible wings that direct the flow to the reactive material. The main issues influencing DART applications are 9 The hydrology of the system, including the degree of heterogeneity. 9 Construction engineering of the DART. 9 Residence time for contaminant removal. The reactive material should be produced in such form that it has a permeability 5 0 200 times greater than that of the surrounding layer. This will direct groundwater to the reactive material from a width of almost twice the well. The design should enable 9 9 9 9 9 9

Maximum open surface area of the DART housing. Maximum groundwater flow rate through the reactive material. Easy production of DARTs with various diameters and lengths. Easy placement of the DART. Easy access to the reactive material. Minimum production costs.

It is estimated that the groundwater residence time in a lO-cm-diameter DART ranges from 1.4 h for an outer flow paths to 29.3 h through the centre of DART and this compared favourably with residence times needed for 99% uranium removal in a metallic iron PRB in Fry Canyon, Utah (Freethy et al., 2002).

44

T. Meggyes

2.7. Applying reactor barrier technology to existing containment systems Although cut-off technology is highly sophisticated and in most cases provides efficient containment, the zone of groundwater separated by a cut-off wall from the surrounding reservoir very often requires treatment which results in additional costs. Applying reactive barrier technology to an existing containment system by opening up the cut-off wall at distinct points and filling the openings with reactive material thus transforming containment into a funnel-and-gate system. In this way, the groundwater can be suitably treated and it need not be pumped, which helps reduce costs (Bradl and Bartl, 1999).

2.8. Injected systems Injecting reactive materials into the ground at high pressure without strict geometrical boundaries in terms of a "wall" is an attractive idea which provides a high degree of flexibility. Nevertheless, it is necessary to ensure that the contaminant plume is efficiently treated and that there is no by-pass flow, which may impair the remediation effect. An injected system can control large and deep plumes even if their extent is irregular (Golder Associates Ltd., 1998). Techniques developed over many decades in petroleum engineering for strata fracturing, secondary recovery, etc. provide extensive knowledge and experience of injection techniques. The reactive material cannot be recovered in injected systems, and therefore degradation principles should be preferred. Two main types of injected systems seem to have so far been developed Injection into existing pores in the ground. 9 Ground fracturing.

9

The effective injection radius largely depends on the pores available and injection wells may require a very close spacing in finely grained soils. Some authors suggest that 10 -5 m/s is the permeability limit below which only pure liquids may be injected. Injecting bacteria, air micro-bubbles as an oxygen supplier (Duba et al., 1996; Koenigsberg, 1998), or cationic surfactants (Burris and Antworth, 1990, 1992) have also been suggested.

2.9. Hydraulic fracturing Ground fracturing, as used in petroleum engineering to enhance permeability around oil wells, aims at creating an improved zone by cracking the ground under high pressures and pumping sand or similar granular material into the cavity created. Using this method within the concept of barrier technology (Murdoch et al., 1997; Gavaskar, 1999), a part of the material filled in may be reactive material intended for pollutant treatment. As in petroleum engineering, the fractures are initiated from wells and are nearly horizontal structures so that they are capable of intercepting downward-moving contaminant plumes. In addition to introducing reactive materials into the ground, another advantage is that a high-permeability zone is formed which encourages groundwater flow. Fractured zones may also be applied to direct groundwater flow towards the gates in funnel-and-gate systems (Golder Associates Ltd., 1998). Murdoch et al. (1997) found that the pressure required to fracture the ground increased with the depth but was surprisingly low: 500 kPa


45

maximum was measured during a test at 2 m depth. The fractures usually have a preferred direction of propagation, they are therefore asymmetric with respect to the borehole and they climb in the preferred direction of propagation. Murdoch et al. (1997) report on 7 10 m fracture dimensions in silty clay, with O.1-1.25 m 3 filled-in material and 5-25 mm fracture thickness. The fracturing procedure can be monitored by recording both pressure and deformation of the ground surface, which will lift to form a gentle dome. Applications include the insertion of materials that alter redox conditions, adsorb contaminants or slowly release useful materials (e.g. oxygen, nutrients, porous ceramic granules, etc.). By using high-energy jets and adjustable outlets, directional fracturing is feasible. Vertical fractures can also be constructed by hydraulic fracturing initiated from a series of wells (Gavaskar, 1999). The fractures are monitored through techniques such as downhole resistivity sensing to ensure coalescence or overlap of the fractures. A 8-1O cm barrier thickness can be achieved and some variability in thickness can be addressed by deploying two parallel barriers.

2.10. Biobarriers Subsurface biofilm barriers or biobarriers consist of polymer films and cells of microorganisms, together with captured organic and inorganic particles (Cunningham et al., 1991, 1997 Sharp and Cunningham, 1998). Biobarriers may selectively plug permeable strata with microbial biomass thus influencing the hydraulic conductivity. Reduction of the hydraulic conductivity by five orders of magnitude is possible. Biobarrier technology may also be useful as a means of funnelling contaminated groundwater through subsurface treatment systems (e.g. funnel-and-gate). The first step in forming a biobarrier is the isolation and identification of potential barrier-forming bacteria from a field site. Once identified, these bacteria will be re-injected to serve as the inoculum for biobarrier formation. Extra-cellular polymer (EPS)-producing bacterial strains (i.e. mucoid phenotype) are desirable candidates for barrier formation. Several pseudomonas and klebsiella strains have been isolated from potential field sites. In addition to exhibiting high EPS production, these bacteria also have the ability to biodegrade benzene, toluene, xylene (BTEX) compounds, thereby raising the possibility that biobarriers can be constructed which actively biodegrade dissolved contaminants in addition to providing containment. These bacteria were found to grow well on either molasses or distillery waste, which means that a low-cost nutrient source is potentially available for field scale biobarrier construction. The next step is to verify that biobarrier formation and hydraulic conductivity reduction can be achieved in laboratory scale packed bed reactors. Biobarrier experiments carried out by Cunningham in 0.9 m long, 15 cm diameter PVC and stainless steel columns under a hydraulic gradient of 1.0 m/m yielded a reduction in the hydraulic conductivity from an initial value of about 4 cm/min down to approximately O.Ol cm/min. Similar experiments were run in 0.3 • 0.9 • 0.15 m 3 stainless steel lysimeters under a head of 0.03 m/m and reductions from 1 cm/min to approximately 10 -5 cm/min were obtained. Biobarriers can be constructed by using an array of wells to introduce micro-organisms, nutrients and oxygen into the ground (Fig. 2.4). The activity of micro-organisms leads to the development of contiguous columns of low permeability which in tum form the barrier.

46

T. Meggyes

Figure 2.4. Schematicbiobarrier configuration (Hiebert, 1998).

Preliminary economic estimates by Hiebert (1998) indicated a cost range of US$ 6.510.7 million for a biobarrier and US$ 9.8-13.5 million for a grout curtain for a 30 m deep and 3200 m long barrier. A sheet-pile wall of the same length but only 12 m depth costs U S $ 1 5 - 1 7 million. Biodegradation of trichloroethylene can be enhanced by combining Burkholderia cepacia PRI-pTOM31c, a TCE-degrading organism unable to form a stable biofilm, with Klebsiella oxytoca, a thick biofilm-forming organism (Koml6s et al., 2001). It is necessary for the bacterial population to produce enough biofilm material to attach to a surface in addition to carrying out the desired reaction. B. cepacia and K. oxytoca can co-exist in a dual-species biofilm and growth rate is not an adequate predictor of which organism would out-compete the other. Rather, substrate concentration is a dominant variable in controlling the population distribution of the two organisms in a biofilm: at high substrate concentration K. oxytoca is the dominant organism, while at lower concentrations B. cepacia becomes the dominant organism. B. cepacia achieves greater population density and higher TCE-degrading potential at the lower substrate level. Varying the substrate concentration can be used to regulate the fraction of each organism in a dual-species environment in biofilms and porous media. Large-scale field experiments are important tools in further developing biobarriers towards practical application. Cunningham et al. (2003) report on a successful long-term experiment at a field-related scale. An engineered microbial biofilm barrier capable of reducing aquifer hydraulic conductivity, while simultaneously biodegrading nitrate, has been developed and tested in a 40 m • 60 m and 6 m deep PVC-lined test cell. A flow field was established across the test cell with an initial hydraulic conductivity of 4.2 • 10 -4 m/s by injecting water upgradient and pumping from an effluent well downgradient. A lO m wide biofilm barrier was developed along the centreline of the test cell by injecting a starved bacteria inoculum of Pseudomonas fluoresces strain CPC2IIa, followed by injection of a growth nutrient mixture composed of molasses, nitrate and other additives.


47

A 99% reduction of average hydraulic conductivity across the barrier was accomplished after 3 months of weekly or bi-weekly injections of growth nutrient. Reduced hydraulic conductivity was maintained by additional nutrient injections at intervals ranging from 3 to l O months. After the barrier was in place, a sustained concentration of 100 mg/l nitrate nitrogen, along with a l O0 mg/l concentration of conservative (chloride) tracer, was added to the test cell influent over a 6-month period. At the test cell effluent the concentration of chloride increased to about 80 mg/l while the effluent nitrate concentration varied between O.O and 6.4 mg/l. Biobarriers may also act as a remediation tool. Spinnler et al. (2004) report on the removal of a methyl tertiary-butyl ether (MTBE) and tertiary-butyl alcohol (TBA) contamination using a biobarrier. In the bioaugmentation process used specialised MTBEdegrading bacteria (MC-1OO) were added to unconsolidated sediments in the saturated zone. Oxygen was pulsed into the bacteria-containing zone to increase the dissolved oxygen content. The combination of ethene-degrading microbes and oxygen created a biobarrier in which MTBE and other gasoline oxygenates were degraded thus limiting the mobility of the contaminant plume.

2.11. Combined PRB - phytoremediation systems Combining reactive barriers with biological techniques may prove efficient for complex pollutants when both physical, chemical and biological effects may be of benefit. For example, phytoremediation and biodegradation can increase clean-up efficacy in PRBs for creosote, a complex agent used for wood preservation and consisting of polycyclic aromatic hydrocarbons (PAHs), phenols and nitrogen/sulphur/oxygen heterocyclic aromatics (Rasmussen, 2002). Rassmussen monitored two pilot-scale barriers installed on an abandoned creosote-contaminated wood preservation site for 29 months. One of the barriers consisted of a soil/sand section that was vegetated with orchard grass (Dactylis glomerata) and a peat/sand section. The function of the first section was to destroy contaminants by bioremediation in a rhizosphere environment and the second section was aimed to sorb or degrade compounds not removed in the first section. The second barrier consisted of a compost/sand mix. The same barrier materials were tested in bench-scale tests in the laboratory. Effluent concentrations of the two field barriers were consistently low during the 29 months of operation, though groundwater temperature was between 0.4 and 13~ Laboratory tests at 9~ indicated that sorption was important for the removal of PAHs and NSO-compounds and biological degradation was dominating for phenols. In the field study the vegetated soil/sand section was least efficient and was partly saturated with PAHs. The low treatment efficiency was probably caused by anaerobic conditions. In the laboratory, the presence of vegetation in soil/sand barrier material improved the treatment of creosote-contaminated groundwater. The effect was most obvious for dimethylphenols and trimethylphenols. The results indicate that biodegradation is an important process in combined PRB/phytoremediation systems, which may efficiently treat creosote-contaminated groundwater. For biological degradation to be effective, aerobic conditions are required. Microbial degradation seems to significantly extend the lifetime of the barrier.

T. Meggyes

48

2.12. Biopolymer trenching Biopolymers were first used as additives to stabilise trench walls during excavation. Through the advancements achieved by their use, the additive application has been developed into a new technology (Day et al., 1999; Gavaskar, 1999; Gavaskar et al., 2000; Sivavec et al., 2002). The procedure exhibits similar features to traditional slurry trenching, the main difference being that biopolymer (e.g. guar gum) is added to the stabilising suspension. The reactive material can then be placed in the trench using tremie pipes. In the construction of a 7.6 m long and 10.4 m deep experimental reactive barrier, 90/10% iron/sand mixture was placed into the trench using a tremie pipe (Sivavec et al., 2002). Care was taken to minimize segregation and contact between the iron/sand mixture and the biopolymer. Following placement high pH enzyme breaker was added to the fluid to break the remaining biopolymer in the trench and clay was placed on the barrier to prevent contact with air.

3. Cost analysis Though it is generally appreciated that costs are usually site-specific, it is as a rule true that increasing depth leads to increasing cost. It is also true that brown-field sites, constituting the majority of the cases, are more expensive than green-field sites because any existing facilities, pipes, cables and other infrastructure will hamper construction. Experience available so far does allow, however, some limited conclusions on costs. A few data are compiled in Table 2.5. 4. Outlook

There are a large range of construction techniques by which PRBs can be deployed for the in situ treatment of contaminated groundwater. Many construction methods used so far rely on techniques developed for cut-off walls. With the development of the reactive barrier technology, deeper and more difficult installations will be undertaken, which will require more advanced construction methods. Interest in the improvement of construction methods has recently increased.

Table 2.5. Construction costs of PRBs in the late 1990s (Gavaskar et al., 2000). Construction technique

Maximum depth (m)

Caisson-based construction Mandrel-based construction Continuous trenching Jet technology Deep soil mixing Hydraulic fracturing Vibrating beam

> 15 12-15 8 60 50 25-40 30

a

Without mobilisation costs, based on 1C = 1.255.

Costsa 120- 700 C/m

80-200 C / m 2 40-100 C / m 2 300-1600 C / m 2 80-200 C / m 3 1800 C per fracture 60 C/m2


49

Acknowledgements The author wishes to acknowledge the following; the Solid and Hazardous Waste Research Unit of the University of Newcastle Upon Tyne, for permission to use parts of the S t a t e - o f - t h e - A r t report "Landfill L i n e r S y s t e m s " ; L a n d C o n t a m i n a t i o n and Reclamation for permission to use parts of " R e m o v a l of organic and inorganic pollutants from groundwater using permeable reactive barriers. Part 2. Engineering of permeable reactive barriers"; and the European Science Foundation for supporting a Workshop on active and passive groundwater remediation technologies whose results were published in a book by Thomas Telford Publishers (Simon et al., 2002a), for permission to use parts of "Advanced Groundwater Remediation".

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Cunningham, A., Warwood, B., Sturmann, P., Horrigan, K., James, G., Costerton, J.W. et al. 1997. Biofilm processes in porous media - practical applications. In: Amy, P.S., Haldeman, D.L. et al. (Eds), The Microbiology of the Terrestrial Deep Surface. Lewis Publishers, Boca Raton, NY, pp. 325-344. Cunningham, A.B., Sharp, R.R., Hiebert, R., James, G., 2003. Subsurface biofilm barriers for the containment and remediation of contaminated groundwater. Bioremed. J. 7 (3-4), 151-164. Dash, J.G., Fu, H.Y., Leger, R., 1997. Frozen soil barriers for hazardous waste confinement, International Containment Technology Conference, St. Petersburg, FA, USA, pp. 607-613. Day, S.R., O'Hannesin, S.F., Marsden, L., 1999. Geotechnical techniques for the construction of reactive barriers. J. Hazard. Mater. B67, 285-297. Debreczeni, E., Meggyes, T., 1999. Construction of cut-off walls and reactive barriers using jet technology. In: Christensen, T.H., Cossu, R., Stegmann, R. (Eds), Sardinia 99, Seventh International Waste Management and Landfill Symposium, S. Margherita di Pula, Sardinia, Italy, CISA Environmental Sanitary Engineering Centre, Cagliari, Italy, Conference Proceedings, Vol. IV, pp. 533-540. Dietrich, J., M~irten, A., Feeser, V., 2004. Chemische Best~indigkeit von Dichtwandmassen gegentiber schadstoffhaltigen sickerw~issern. 11. Braunschweiger Deponie- und Dichtwandseminar 2004. Mitteilung des Instituts ftir Grundbau und Bodenmechanik, VoI. 74. Technische Universit~it Braunschweig, pp. 335-353, in German. Duba, A.G., Jackson, K.L., Jovanovich, M.C., Knapp, R.B., and Taylor, R.T., 1996. TCE Remediation using in situ resting state bioaugmentation. Environmental Science and Technology 39(6), 1982-1989. Dtillman, H., Geil, M., Zirfas, J., 1993. Einkapselung von Sonderabfalldeponien. In: Jessberger, H.L. (Ed.), Sicherung von Altlasten. Rotterdam: Balkema. Dtillman, H., 1999. Geotechnical Bureau Prof. Dr.-lng. H. Dtillmann, Aachen, Germany, statement on cutoff walls. Sardinia 99, Seventh International Waste Management and Landfill Symposium. Margherita di Pula, Sardinia, Italy. Dwyer, B., 1998. Remediation of deep soil and groundwater contamination using jet grouting and innovative materials, Subsurface Barrier Technologies Conference, Tucson, AZ, USA, International Business Communications. Fischer, J., 2002. Dichtw~inde gegen hohes Schadstoffpotential. Entwicklung, Einsatz, Aussichten. 10. Braunschweiger Deponieseminar 2002, Mitteilung des Instituts fur Grundbau und Bodenmechanik, VoI. 69. Technische Universit~it Braunschweig, pp. 337-352, in German. Freethey, G.W., Naftz, D.L., Rowland, R.C., Davis, J.A., 2002. Deep aquifer remediation tools: theory, design, and performance modeling. In: Naftz, D.L., Morrison, S.J., Davis, J.A., Fuller, C.C. (Eds), Handbook of Groundwater Remediation Using Permeable Reactive Barriers. Academic Press, San Diego, CA, pp. 133-161. Gavaskar, A.R., 1999. Design and construction techniques for permeable reactive barriers. J. Hazard. Mater. 68, 41-71. Gavaskar, A., Gupta, N., Sass, B., Janosy, R., Hicks, J., 2000. Design Guidance for Application of Permeable Reactive Barriers for Groundwater Remediation. Final Report, Battelle, Columbus, Ohio. German Geotechnical Society, 1993. Geotechnics of Landfills and Contaminated Land: Technical Recommendations "GLC" for the International Society of Soil Mechanics and Foundation Engineering. Verlag Ernst und Sohn, Berlin. Ghezzi, G., Ghezzi, P., Pellegrini, M., 1999. Use of a cement-bentonite-slurry plastic diaphragm with hdpe membrane for MSW landfill. In: Christensen, T.H., Cossu, R., Stegmann, R. (Eds), Sardinia 99, Seventh International Waste Management and Landfill Symposium, S. Margherita di Pula, Sardinia, Italy, CISA Environmental Sanitary Engineering Centre, Cagliari, Italy, Conference Proceedings, Vol. IV, pp. 549-554. Golder Associates Ltd. 1998. Active Containment: Combined Treatment and Containment Systems. Department of the Environment, Transport and the Regions, ISBN 185112 114 5, London. Gravelding, D., 1998. Design and construction of a 1200 ft funnel and gate system, Subsurface Barrier Technology Conference, Tucson, AZ, USA, International Business Communications. Hiebert, R., 1998. Using biological barriers to control movement of contaminated groundwater, Subsurface Barrier Technologies Conference, Tucson, AZ, USA, International Business Communications. Hitze, R., 1987. Einsatz der schadstoffresistenten Dynagrout-Abdichtungsmaterialien, geschildert an konkreten Beispielen der Schadstoff-Umschliel3ung durch Schlitzwandbau und Injektionstechnik. In: Seminar tiber Altlasten und kontaminierte Standorte. Ruhr-Universit~it Bochum.


51

Hubble, D.W., Gillham, R.W., Cherry, J.A., 1997. Emplacement of zero-valent metal for remediation of deep contaminant plumes, International Containment Technology Conference, St. Petersburg, FA, USA, pp. 872-878. Jansen, T., Grooterhorst, A., 1999. Reaktive Schmalw/inde zur passiven Grundwasserreinigung. TerraTech 3, 46-48, in German. Jessberger, H.L., 1992. ITVA Arbeitspapier Dichtw/inde. Jessberger, H.L., Geil, M., 1991. Eigenschaften und Anforderungen an Dichtwandmassen. In: Franzius, V., Stegmann, R., Wolf, K., Brandt, E. (Eds.), Handbuch der Altlastensanierung. Heidelberg: R.v. Decker' s. 5.3.3.0.2. Jessberger, H.L., Geil, M., 1992. Einsatz von Spundw/inden bei Deponien und Altlasten. Geotechnik (4), 237-242. Kellner, C., Scheibel, B., 2004. Innovative SicherungsmaBnahmen ftir Deponien-Ausftihrung der Kammerdichtwand Vorketzin. 11. Braunschweiger Deponie- und Dichtwandseminar 2004. Mitteilung des Instituts ftir Grundbau und Bodenmechanik, VoI. 74. Technische Universit/it Braunschweig, pp. 363-380, in German. Koenigsberg, S., 1998. The formation of oxygen barriers with ORC. Subsurface Barrier Technologies Conference, Tucson, Arizona, USA. Koml6s, J., Cunningham, A., Camper, A., 2001. Varying substrate concentration to enhance TCE degradation in dual-species bioreactors. In: Magar, V.S., Fennel, D.E., Morse, J.J., Alleman, B.C., Leeson, A. (Eds), Anaerobic Degradation of Chlorinated Solvents, The Sixth International In-Situ and On-Site Bioremediation Symposium, San Diego, California, June 4-7. Battelle Press, Columbus, Richland, pp. 117-124. Kutzner, C., 1991. Injektionen im Baugrund. Enke Verlag, Stuttgart. Mageau, D., 1998. Use of a frozen ground barrier to contain groundwater contamination, Subsurface Barrier Technologies Conference, Tucson, AZ, USA, International Business Communications. Meggyes, T., Pye, N., 1995. Landfill Liner Systems, N1-28, Federal Institute of Materials Research and Testing (BAM), Berlin. In: Holzl6hner, U., August, H., Meggyes, T., Brune, M. (Eds), Solid and Hazardous Waste Research Unit. The University of Newcastle, Penshaw Press, Sunderland, UK. Meggyes, T., Simon, F.-G., 2000. Removal of organic and inorganic pollutants from groundwater using permeable reactive barriers. Part 2. Engineering of permeable reactive barriers. Land Contam. Reclamation 8 (3), 175-187. Meseck, H., 1987. Dichtw/inde mit eingestellten Kunststoffbahnen (Kombinationsdichtw~inde). In: Meseck, H. (Ed.), Dichtw/inde und Dichtsohlen, Fachseminar 02/03. Juni, 1987, Mitteilung des Instituts ftir Grundbau und Bodenmechanik, VoI. 23. Technische Universit/it, Braunschweig, pp. 155-170. Morrison, S.J., Carpenter, C.E., Metzler, D.E., Bartlett, T.R., Morris, S.A., 2002. Design and performance of a permeable reactive barrier for containment of uranium, arsenic, selenium, vanadium, molybdenum and nitrate at monticello, Utah. In: Naftz, D.L., Morrison, S.J., Davis, J.A., Fuller, C.C. (Eds), Handbook of Groundwater Remediation Using Permeable Reactive Barriers. Academic Press, San Diego, CA, pp. 371-399. Murdoch, L., Slack, B., Siegrist, B., Vesper, S., Meiggs, T., 1997. Advanced hydraulic fracturing methods to create in situ reactive barriers, International Containment Technology Conference, St. Petersburg, FA, USA, pp. 445-451. Puls, R.W., 2001. Personal Communication. Rasmussen, G., 2002. Sorption and Biodegradation of Creosote Compounds in Permeable Barriers. Doctor Scientiarium Thesis, Agricultural University of Norway, ,~s. Rodatz, W., 1994. Grundbau, Bodenmechanik, Unterirdisches Bauen. In: Institut fur Grundbau und Bodenmechanik. Braunschweig, in German. Roth, S., 1988. Eignung von Stahlspundw~inden ftir Einkapselung von Altlasten. In: Handbuch der Altlastensanierung. R.v. Deckers, Heidelberg, in German. Schulze, B., 1992. Injektionssohlen-Theoretische und experimentelle Untersuchungen zur Erh6hung der Zuverl~issigkeit. Institut ftir Bodenmechanik und Felsmechanik der Universit~it Karlsruhe, Report, No. 126 (in German). Schultze, B., Mu6otter, T., 2001. Sanierung des ehemaligen Gaswerksgel~indes Karlsruhe-Ost mit funneland-gate (Aktivkohle). In: Burkhardt, G., Egloffstein, T., Czurda, K. (Eds), ALTLASTEN 2001 Neue Verfahren zur Sicherung und Sanierung, Conference Proceedings, VoI. 4. ICP Eigenverlag Bauen und Umwelt, Karlsruhe, pp. 47-58, in German.

52

T. Meggyes

Sharp, R.R., Cunningham, A., 1998. Fundamentals of biobarriers, design, development and activity, Subsurface Barrier Technology Conference, Tucson, AZ, USA, International Business Communications. Seitz, J., 1987. Zementfreie dichtwandmasse. Gro6versuch zur baupraktischen Anwendung. BaustoffRecycling 5, 2-3. Simon, F.-G., Meggyes, T., McDonald, C., (Eds), 2002a. Advanced Groundwater Remediation - Active and Passive Technologies. Thomas Telford, London. Simon, F.-G., Meggyes, T., Ttinnermeier, T., 2002b. Groundwater remediation using active and passive processes. In: Simon, F.-G., Meggyes, T., McDonald, C. (Eds), Advanced Groundwater Remediation Active and Passive Technologies. Thomas Telford, London, pp. 3-34. Sivavec, T., Krug, T., Berry-Spark, K., Focht, R., 2002. Performance monitoring of a permeable reactive barrier at the Somersworth, NH Landfill Superfund Site. In: Simon, F.G., Meggyes, T., McDonald, C. (Eds), Advanced Groundwater Remediation - Active and Passive Technologies. Thomas Telford, London, pp. 87-100. Smyth, D.A., Shikaze, S.G., Cherry, J.A., 1997. Hydraulic performance of permeable barriers for the in situ treatment of contaminated groundwater. Land Contam. Reclamation 5 (3), 131-137. Spinnler, G.E., Salanitro, J.P., Manner, P.M., 2004. In situ remediation of groundwater contaminated with MTBE and TBA using a biobarrier, First International Symposium on Permeable Reactive Barriers, PRB-Net 2004, Belfast 14-16 March, Abstracts Proceedings. Stahlmann, J., Scholz, Chr., 2004. Ursachen und Auswirkungen von Imperfektionen bei Dichtw~inden. 11. Braunschweiger Deponie- und Dichtwandseminar 2004. Mitteilung des Instituts ftir Grundbau und Bodenmechanik, 74. Technische Universit~it Braunschweig, pp. 363-380, in German. Stroh, T., Sasse, D., 1987. Beispiele fiir die Herstellung von Dichtw~inden im Schlitzverfahren. In: Meseck, H. (Ed.), Dichtw~inde und Dichtsohlen, Fachseminar 02/03. Juni 1987, Mitteilung des Instituts ftir Grundbau und Bodenmechanik, VoI. 23. Technische Universit~it Braunschweig, pp. 35-38, in German. T6th, S., 1989. Soilcrete-Dichtwand als vorsorgliche Sicherung zur Verhinderung der Ausbreitung von Kontaminationen. Geotechnik 1, 1-4. Weber, H.H., Fresenius, W., Matthess, G., Mtiller-Kirchbauer, H., Stow, K., Wel3ling, E., 1990. Altlasten: Erkennen, Bewerten, Sanieren. Springer, Berlin. Wilson, R.D., Mackay, D.M., 1997. Arrays of unpumped wells: an alternative to permeable walls for in situ treatment, Proceedings of International Containment Technology Conference, St. Petersburg, Florida, USA, pp. 888-894.


53

Chapter 3

Materials and processes for uranium removal from contaminated water Dimitris Panias, Anthimos Xenidis and Athina Krestou

A. Introduction Cationic metals usually have limited mobility in soils and aquifers, particularly in those with high clay and organic matter contents, high alkalinity and low hydraulic permeability (Fetter, 1993). However, complexing agents such as carbonates, hydroxides, sulfates, phosphates, fluorides and possibly silicates which are present in natural waters tend to increase the solubility of metals (Langmuir, 1978). Some metals and metalloids form anions or oxyanions with high biogeochemical mobility and therefore pose considerable environmental risk. Examples include chromium (chromate), selenium (selenate), arsenic (arsenate) and molybdenum (molybdenate). To date most Permeable Reactive Barrier (PRB) projects have focused on organic contaminants such as chlorinated hydrocarbons, however, the use of PRBs to treat inorganic contaminants is of great interest especially for mine water treatment (including the treatment of acid mine drainage) and the remediation of radionuclide contaminants. Chromate and uranium (as the uranyl cation) were among the first inorganic compounds studied in relation to their treatment by PRBs. An overview of the treatment of inorganic contaminants using PRB is given by Blowes et al. (2000) and Naftz et al. (2002). The engineering design of a PRB involves the selection of a reactive material that will reduce the concentration of the target contaminant to below the desired groundwater concentration for the design-life of the barrier without releasing any harmful substances back into the water. The only way to be certain that a particular reactive material will achieve this objective is to understand the attenuation processes by which it operates. This chapter describes the performance of selected reactive materials at removing uranium from aqueous solutions, and discusses the attenuation processes associated with each.

B. Materials and experimental procedures The effectiveness of natural zeolitic tuff, hydroxyapatite (HAP), activated carbon (AC), hydrated lime (Ca(OH)2) and elemental iron (Fe o, usually referred to as "zero-valent iron") as reactive materials for the removal of uranium from groundwater has been investigated. The selection of these materials was based on PRB studies reported in the literature (Simon and Meggyes, 2000).

54

D. Panias, A. Xenidis, A. Krestou

Table 3.1. Selected physical and chemical properties of the studied reactive materials.

Parameter

Natural zeolitic tuff

Hydroxyapatite

Activated carbon

Ca(OH)2 Elemental iron

Particle size (mm)a Density (kg/dm3)a Specific surface area (m2/g)a Paste pHb

0-1.2 2.16 15.6

0-1.25 0.5 65

1.5-4 0.48 1000

0-0.59 NA NA

0.35-1.2 6.7-7.2 0.048

7.9

6.5

10.2

12.4

5.1-5.3

NA, not analysed. a Data given by supplier of the material. b EPA Method 9045c (measured in water suspension, water/solid = l/l). Selected physical and chemical properties of the studied materials are presented in Table 3.1. The sample of natural zeolitic tuff was supplied by Silver & Baryte Mining Co. and originates from Pentalofos, Greece. It was crushed and milled upon delivery. The mineralogical composition of the natural zeolite is comprised primarily of clinoptilolite (--~ 85%), feldspar (5%), montmorillonite (4%) and quartz. The cation exchange capacity (CEC) of the material was at least 150 meq/10O g measured with the ammonium acetate method. The hydroxyapatite sample was provided by Chemische Fabrik Budenheim, Germany. Chemical analysis of the HAP sample showed that its main constituents are CaO (58.2%), P205 (41%) and MgO (0.43%), while K20, Fe203 and MnO are present in lower quantities. X-ray diffraction (XRD) mineralogical analysis showed the existence of two types of hydroxyapatite, with Cas(PO4)3OH being the main mineralogical phase, while Cas(PO4)3(OH, CI, F) is of minor importance. The activated carbon sample was provided by Donau Chemie, Austria, with the trade name DonauCarbon | CC15, and was supplied as cylindrical pellets. The hydrated lime sample was supplied in industrial grade by Mosholios SA, Greece. The elemental iron sample was supplied by Gotthart Maier, Germany and is characterised as cast iron grit. Chemical analysis showed that the elemental iron consisted of Fe (92.0%), C (3.3%), Si (2.0%) and other elements such as Mn, AI, S, Ni, Cr and P in concentrations lower than 1%. Mineralogical investigation carried out by XRD and scanning electron microscopy revealed the existence of metallic iron and some graphite and iron oxide. Batch experiments were performed on the materials described above to study their performance at removing of uranium from aqueous solution. Most experiments were carried out using a uranyl nitrate solution with a concentration of lOO0 txg/l uranium and a pH of approximately 1.2. The uranium solutions were prepared by dissolving an appropriate quantity of uranyl nitrate hexahydrate (UO2(NO3)2-6H20), in a diluted solution of nitric acid. The experiments were conducted in agitated glass reactors. The pH of the batch solutions were adjusted to the desired value using NaOH. At the end of each experiment the pulp was filtered and the uranium remained in solution was analysed with the Arsenazo-III spectrophotometric method on a HITACHI U-1100 U V - V i s spectrophotometer.

Materials and processes for uranium removal from contaminated water

55

C. Attenuation processes 1. Zeolites

Zeolites are tectosilicates with three-dimensional aluminosilicate structure containing water molecules, alkali and alkaline earth metals in their structural framework (Gottardi and Galli, 1985). The ability of zeolites to exchange cations in aqueous solution is very well documented. The unbalanced substitution of Si4+ by AI 3+ in the crystal lattice leads to a net negative charge and, subsequently, to the high CEC of most natural zeolites. The selectivity of certain zeolite minerals for specific chemical compounds is defined by pore size and charge properties of the zeolite structure. Zeolites have been widely used as molecular sieves, ion-exchangers, adsorbents, catalysts, detergent builders, etc. Clinoptilolite, a natural zeolite, is a potential material for the remediation of aqueous solutions since it demonstrates a strong affinity for toxic and problematic heavy metals (Loizidou and Townsend, 1987) and can selectively adsorb some radionuclides (Lepperd, 1990). The present study focused on the mechanisms of uranium removal from aqueous solutions by natural zeolitic tuff. Results from batch experiments conducted with uranium solutions containing 1000 Ixg/l of uranium dissolved in deionised water are presented in Figure 3.1. The chemical processes taking place in the solution and on the solution/zeolite interface depend strongly on the uranium species present in solution. To highlight this effect, the uranium speciation in aqueous solution in equilibrium with atmospheric carbon dioxide at 25~ is displayed in Figure 3.2. The diagram was created using the Visual Minteq geochemical modelling software (Gustafsson, 2003). Uranyl removal from aqueous solution showed a strong pH dependency with significant uranyl removal occurring only when the pH exceeded 5 (Fig. 3.1a). The amount of uranyl removed increased as the liquid-to-solid ratio decreases (Fig. 3. lb), despite the CEC of the zeolite exceeding the amount of uranyl available in the tests by a factor of at least 40. As shown in Figure 3.1, uranium is not removed from the solution at pH < 3, where uranyl ions (UO 2+) are the dominant species in aqueous solution (Fig. 3.2). As pH increases from 3 to 5, uranium removal increases substantially reaching a maximum value at pH very close to 5. In the same pH range the stability of the mononuclear and polynuclear hydroxo-uranyl complex ions UO2OH + and (UO2)2(OH) 2+ in the solution increases gradually. Therefore, the uranyl ions are not the only uranium species in the solution although they comprise the majority of them. These changes in the composition of the aqueous solution alone cannot explain the experimental results and an explanation has to be sought based on the physicochemical properties of the solution/zeolite interface. In aqueous solutions the zeolite surface has a permanent negative charge (Fig. 3.3), mainly due to the isomorphous substitution of some Si 4+ by AI 3+ within the crystal lattice. These sites are located in the siloxane layer and comprise the main sites accessible to ion exchange reactions with various cations, but they are located inside the zeolite channels and are not accessible to uranium species. The estimated size of a hydrated uranyl cation [UO2(H2O)5] 2+ is around 6.5 A (Krestou et al., 2003) and is therefore much greater than the mean dimension of the zeolite channels (about 5 A), indicating that the U(VI)-bearing species in the solution have no access to the exchangeable sites of the zeolitic tuff. o


56 (a) 100 ~,

41'

80-

v m >

o E

60-

E

40-

t-l,...

D

200 0

i

i

!

!

4

6 pH

8

10

12

(b) 100 90v m

o E

80-

E

70-

:D

60-

.m C t~

50

i

O

1

I

i

2

3

4

Zeolite concentration (g/I)

Figure 3.1. (a) U(VI) removal by zeolite as a function of pH of the uranyl nitrate solution (solid/liquid ratio 2 g/l, contact time 2 h, ambient temperature). (b) Effect of the solid/liquid ratio on the removal of U(VI) (at pH 5, contact time 1 h, ambient temperature). Therefore, cation exchange is highly unlikely to be the major mechanism of uranium (VI) removal from water by such zeolite minerals. In addition to the exchangeable sites located in the zeolite channels, there are structural hydroxyls like silanol ( S i - O H ) and aluminol ( A I - O H ) groups located on the external surface of zeolites. These groups are directly accessible to potential determining ions and uranium species in aqueous solutions. The structural hydroxyl groups (depicted below as SOH) show an ampholytic behaviour and can react as an acid or a base - depending on the solution pH: SOH ~ + H + ,--, SOH~-

(in acidic solutions)

SOH ~ + O H - ,---, S O - + H 2 0

(in alkaline solutions)

(3.1) (3.2)

Therefore, depending on the solution pH, positively or negatively charged local sites exist on the external surface of zeolites where uranium adsorption can take place.

Materials and processes for uranium removal from contaminated water 100% 80%

v

m

60%

E E

40%

tO

tL~

E)

:U07+~"~

(UO2)3(OH)~

.................................................

. . . . . . . . . . . . . . . . . . . . . . . . . . . .

-I.... i-I

20% ................

0%

.....

C~ 3(aq)

......... /---l-

2

0

57

4

6 pH

8

10

12

Figure 3.2. Uraniumspeciation in an aqueous solution at equilibrium with atmospheric carbon dioxide and total uranium concentration equal to 5 x 10-6M (1190 I~g/l).

An estimation of this variable surface charge determined by acid-base titration is presented in Figure 3.4, where the surface charge of the natural zeolitic tuff is plotted as a function of solution pH. The acid-base titration has been performed in the presence of 1 M NaNO3 in solution, so that Na + ions comprise the main exchangeable ions. Therefore, the difference between the concentration of protons and hydroxide ions can be attributed almost exclusively to reactions taking place on the particle surface. The specific surface charge of zeolite decreases steeply as the pH increases from 4 to 5.5 (Fig. 3.4). As a result, the work necessary to move a positively charged uranyl ion to the positively charged surface of the zeolite has substantially decreased and is at least six times lower at pH 5.5 than at pH 4, and uranium adsorption by the zeolite is enhanced as

-10 -15 :~ -20

.....

-~- 9

v m

.~- -25

9149

t-

(l.)

9

o - 30

_ 9

13.

"~ -35 N

-i

II

-40 -45

o

i

i

4

i

i

6 8 pH (final)

i

10

i i

12

14

Figure 3.3. Zeta potential of clinoptilolite as a function of pH (after Onal et al., 2001).

58

D. Panias, A. Xenidis, A. Krestou 1,5 (.9 v

03 tO

1,00,5-

.... O ....

0,0-

O==-==O . . . . .

O

~

(D

o -0,5 =

-1,0

-~

-1,5 -2,0 -

-2,5 4

I

I

I

i

I

5

6

7

8

9

10

pH Figure 3.4. Specific surface charge of zeolite as a function of the solution pH (background electrolyte:

1 M NaNO3) (Krestou et al., 2003). the solution pH increases due to this physicochemical change on the solution/zeolite interface. This observation is in accordance with the experimental results shown in Figure 3.1a. The U(VI)-bearing species are preferably adsorbed on the variable charge sites of the silanol groups at the zeolite surface, which are not protonated in this pH region, rather than on the aluminol groups, which are strongly protonated in the same pH region (Krestou et al., 2003). Between pH 5.5 and 9 uranium removal remains high (Fig. 3.la) while the uraniumbeating species in the solution are gradually transformed from the uncomplexed uranyl ions to mononuclear hydroxo-complexes, then to polynuclear hydroxo-complexes and finally to carbonato-complexes (see Fig. 3.2). Moreover, the changes in the ionic composition of the solution are accompanied by alterations in the charge of the uraniumbeating species. Consequently, the charge of the uranium-bearing species is altered from exclusively positive at pH 5 to exclusively negative at pH 8.5. On the other hand, the specific surface charge of zeolite remains very low in the same pH region and for this reason the Coulombic interactions are attenuated. As a conclusion, the size of the polynuclear hydroxo-complexes, the alteration of the charge of uranium beating complex ions and the very low specific surface charge of zeolite are the key factors determining the uranium removal in the pH range of 5-9. At pH values higher than 8.5, uranium removal decreases substantially (Fig. 3. l). In the solution, uranium occurs at such elevated pH values only in the form of carbonatocomplex ions which are highly negatively charged. Moreover, under the same pH values, the U02(C03)34- complex ion prevails in the solution instead of the U02(C03) 2- complex (Fig. 3.2), the former having a more negative charge. Also, the zeolite-specific surface charge is clearly negative at pH higher than 9 (Fig. 3.4), and a further increase of pH moves the surface charge towards even more negative values. As a result, strong repulsive forces are expected between the external surface sites of the zeolite and the uranium species present in the solution, explaining the decreased removal rates of uranium in highly alkaline solutions.


59

Figure 3.5. Effect of the presence of carbonates and sulfates on the uranium removal effectiveness of zeolite (pH 5, solid/liquid ratio 2 g/l, contact time 2 h, ambient temperature). Summarising, the experimental results indicate that the application of zeolites for the removal of hexavalent uranium from aqueous solutions is not very successful. The explanation is thought to be that the uranyl cation is unable to enter the zeolite channels (which are about 5 A in diameter) where the major number of permanent charge sites is located, and the observed uranyl removal was the result of adsorption to the external zeolite surfaces. To study the influence of typical groundwater constituents on the uranium sorption to zeolite, further batch experiments were performed under the same conditions but with the addition of 40Omg/l of dissolved carbonate and 2OOmg/l of dissolved sulfate, respectively. The experimental results shown in Figure 3.5 indicate that the presence of carbonates in the solution causes a significant decrease in uranium removal while sulfate exhibits little influence. As displayed in Figure 3.6 the presence of 400 mg/l of carbonates in the solution totally changes the aqueous uranium speciation. The hydroxo-uranium complex ions disappear from the solution and uranium occurs only in the form of carbonato-complexes. In addition to the changes resulting in the solution composition, alterations on the surface reactivity of the zeolite are also observed. The active sites on the external surface of zeolite, when no carbonates are present, are estimated to be 2.5 x 1019 sites/g while in the presence of 400 mg/l of carbonates they are estimated to be 1.3 x 1019 sites/g (Krestou et al., 2003). This reduction in the number of active sites is equivalent to a decrease of about 50% of the zeolite external surface sites available for uranium sorption. Therefore, uranium removal by sorption on zeolites is negatively affected by the presence of carbonates, which basically rules out the use of zeolites for the remediation of uranium-contaminated groundwater (Krestou et al., 2002).

2. Hydroxyapatite The effect of solution pH, contact time, solid/liquid ratio and the presence of carbonates and sulfates on the uranium (VI) removal by hydroxyapatite from a solution containing

60

D. Panias, A. Xenidis, A. Krestou 100%

" 9

/

UO2(CO3)

3-

80%

iiiiiiiiiiii i-........ ...............i-i!--iiiiiiiiiiiiiiiiiiiiiiiiiiii_

v ,m O

60%

-

40%

-

{3.

E .=2 t--

UOzCO3(aq)

.......................

uo2(co3)~-

t~

20%

-

0%, 2

'

.......i ............ ~

.v.

4

..-.

,.i-

-..

6

..i..

8

..-

i ................

.~

10

12

pH Figure 3.6. Uranium speciation (U concentration: 5 • I0-6M) in the presence of 400mg/l of carbonates.

1000 Ixg/l ( = 4.2 x 10 -6 mol/1) uranium (VI) is displayed in Figure 3.7. The uranium (VI) attenuation by HAP is very high (approx. 95%), occurs rapidly, and is almost independent of the pH, HAP concentration, and the presence of common groundwater constituents such as dissolved carbonate and sulfate. The uranium removal by HAP in the test with a solid/liquid ratio of 0.1 g HAP/I is almost 10 mg U/g HAP. Sequential sorption tests investigating uranyl sorption on HAP indicated that the HAP removal capacity is even higher and at least 20 mg/g. For a better understanding of the phenomena taking place during uranium (VI) removal by HAP, it is necessary to examine the HAP dissolution along with the uranium (VI) chemistry in the solution (Krestou et al., 2004). 2.1. HAP dissolution

The dissolution of HAP in aqueous solutions is described by the following reaction, yielding a caa+/HPO 2- ratio in solution of 5/3 = 1.66: Cas(PO4)3OH(solid) 4-- 4H + *--*5Ca 2+ -I- 3HPO 2- + H2O

(3.3)

The solubility of HAP in water at different pH conditions is displayed in Figure 3.8. The lines represent the total dissolved calcium and phosphate and were calculated based on Reaction (3.3) with Visual Minteq (Gustafsson, 2003), and have been published by Leyva et al. (200 l). It was found in this work that the differences in calculated and measured Ca concentration were consistently small. However, from a pH value of 6.6 upwards the phosphate concentrations were always larger than 1 x 10 -3 mol/l demonstrating an incongruent dissolution of HAP most probably arising from carbonate impurities in

Materials and processes for uranium removal from contaminated water (a) Effect of pH variation (contact time 2 h)

(b) Effect of contact time (pH 7.6)

100 O

100

95

> 95

E

90 E .=_ 85t--

O

4~

~ 9o

__o.-

E ._~ 85

l,,..

D 80

61

;

2

pH

8

D

1'0

80

1'0

9'0

Time (h)

30

(c) Effect of the solid/liquid ratio (pH 8.5, contact time 1 h) 9 ~lOO > O

E

95 9

90

E ._= 85 t-" D 80

O

HAP concentration (g/I)

Figure 3.7. Effect of various parameters on the U(VI) attenuation by hydroxyapatite. All experiments were conducted at ambient temperature and a solid/liquid ratio of 2 g/l (except from c, where the solid/liquid ratio was varied), with a solution containing 1OOOtxg/l of uranium (VI).

1E+00 1E-01 .... : : ~ o 1E-02 ~ E "O 1E-03 t~

~ " ' ~

1E-04

met i

" " " " ~ ~ ~ "~, :

"1E-05 O o 1E-O6 _ 1E-07

3

....

.A_

-O--

"~- .

.

.

.... ~ .. . "

_

o

Bm

.

.

.

OO "~_o

4

;

;

pH

;

8

;

10

Figure 3.8. Calculated solubility of HAP in water at different pH conditions at 2O~ and O.Ol M ionic strength (NaCI). The lines represent the total dissolved calcium (solid line) and phosphate (dashed line). The open circles and squares show measured values for Ca and phosphate for equilibrated HAP samples (after Leyva et al., 2001). Own phosphate concentration measurements are displayed with closed squares and closed triangle (in the presence of 400 mg/l C02-).


62

hydroxyapatite. Leyva et al. (2001) mentioned that the Ca/P ratio in solutions with pH > 6.6 is much lower than 1.66, which corresponds to the theoretical value based on Reaction (3.3), as well as that the dissolution of HAP takes place with simultaneous Ca enrichment of HAP. They attribute this observation to the presence of another reaction taking place in addition to the dissolution process described in Reaction (3.3), in which carbonates partially replace phosphates in the HAP with the phosphates passing into the solution (Regnier et al., 1994).

2.2. Physicochemical properties of the HAP~water interface The HAP/water interface can be described by a chemical model taking into account two different types of surface groups (Wu et al., 1991)" positively charged Ca-OHm- surface species and negatively charged P 0 4 surface species: \ --

__\ C a O H ~

Ca

/

/

\ --Ca /

I n aqueous solutions the a c i d - b a s e

" 0 \p~.O ~

o" \ O-

reactions that are responsible f o r the surface

properties of HAP are as follows =PO-

+ H + ,--, =POH ~

with log K = 6.68

-Ca_OH + ~ -Ca_OH ~ + H+

(3.4)

with log K -- - 9 . 6 6

(3.5)

The surface speciation diagram of HAP in aqueous solutions is presented in Figure 3.9. Negatively charged P O - sites and neutral C a - O H ~ sites predominate in alkaline solutions (Fig. 3.9). Therefore, the surface charge of HAP in alkaline solutions is negative. On the 100%

80% "0

-

60% -

O

u_ 40% -

20% -

---

CaOH~

0% 4

5

6

7

8

9

10

11

12

pH

Figure 3.9. Surface complexes as a function of pH (calculated with equilibrium constants K, see Equations (3.4) and (3.5)).


63

30 2O v

-9

0

C (1)

o -10

Q.

-20

U

-30 I

I

I

I

I

6

7

8

9

10

-40 5

11

pH

Figure 3.10. Calculated zeta potential values versus pH of equilibrated suspensions of synthetic HAP (after Leyva et al., 2001). other hand, positively charged CaOH + species and neutral POH ~ sites prevail in acidic solutions, rendering the surface charge of HAP in this pH range positive. Actually, the zeta potential of hydroxyapatite (Fig. 3.10) is negative when the pH exceeds 7.7, which is considered as the point of zero charge in the HAP/water system. This point of zero charge is shifted to 7.13 when the solution is in equilibrium with atmospheric carbon dioxide (Wu et al., 1991).

2.3. Uranium (VI) solution chemistry Figure 3.11 shows the speciation of uranium in a solution containing 1000 l~g/l uranium in equilibrium with atmospheric C02 and solid (UO2)3(PO4)2. At pH values greater than 100%

u%(c%)]-

......................

80%

....

if)

-O9 60% c~ E .2

40%

_~u%)_~_o_H!~ +_ ..............

_

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

,

C

:D 20%

..........

_,/_

~ UO2H2PO~4

/ uo~Po~

.....

'

'~

0% 2

3

4

5

6

7

8

9

pH

Figure 3.11. Speciationdiagram of uranium species in a solution of 1000 Ixg/l uranium at equilibrium with atmospheric C02 and (UOz)3(PO4)2.

64

D. Panias, A. Xenidis, A. Krestou 8

~ 0

6

E v

4

~

2

-~ 0

0

---O-- Precipitation as uranyl phosphate - - I - - Precipitation as autunite

ffl

E -2 "E -4

~

-6

o

-8

-10

2

i

i

I

I

4

6

8

10

pH

12

Figure 3.12. Uranium solubility in a solution saturated in atmospheric CO2 containing 10-3M total phosphates and 1.6 x 10-3 M total calcium.

7.5 and lower than 3.5, uranium complexes with phosphates and carbonates clearly predominate (Fig. 3.11). Between 5.5 and 7.5 the predominant species are polynuclear hydroxo-complexes while the uranyl ion is the major uranium species in the pH range of 3.5-5.5. The solubility of uranyl phosphate (UO2)3(PO4)2 and calcium uranyl phosphate (Ca(UO2)2(PO4)2, autunite) as a function of pH in a solution containing 1 0 - 3 M total phosphates and 1.6 x 1 0 - 3 M total calcium in equilibrium with atmospheric CO2 is displayed in Figure 3.12. The solubility of uranium is lower in the case of Ca(UO2)2(PO4)2 (Ksp = 10 -47"28, Brown et al., 1981) than in the case of (UO2)3(PO4)2 ( K s p - 10 -49"09, Brown et al., 1981). Moreover, uranium solubility increases substantially in high acidic and alkaline solutions. This is attributed to the uranium complexation with phosphates in the acidic area and with phosphates and carbonates in the alkaline area (Fig. 3.11). In addition, Figure 3.12 shows that the initial uranium content of the solution (4.2 x 10 -6 mol/l) exceeds the solubility of autunite in the pH range between 3 and 9, and the solubility of uranyl phosphate in the pH range between 3.5 and 8. Therefore, uranium can theoretically be removed from an aqueous solution containing 1 0 - 3 M total phosphates between the mildly acidic and the moderate alkaline pH by a precipitation mechanism. Although uranium precipitation in the form of phosphate salts is theoretically not feasible for pH values higher than 9, it is noteworthy to mention the differential behaviour of calcium as a function of pH in a solution containing 10 -3 M total phosphates (Fig. 3.13). For a pH higher than 9, calcium forms insoluble carbonate minerals rather than phosphate minerals, and thus calcium is removed as CaCO3. This change in the mechanism of calcium removal probably affects the mechanism of uranium removal. Therefore, it can be deduced that uranium can be precipitated in the form of calcium dioxouranium (VI) carbonate (CAU02(C03)2) when the pH is higher than 9. Although this precipitate is a known insoluble compound in the literature (Grenthe et al., 1992), there is a lack of thermodynamic data concerning its formation. In addition to the formation of this


65

10

~"

8-

E

6-

.Q

4-

O

_= O (n E .m (3

- - i - - Calcium carbonate

0-

rj

-2-

o

-4-

E~

---- Calcium phosphate

2-

-6

2

~lt

I

6

I

pH

8

i

10

12

Figure 3.13. Calcium solubility as a function of pH in a solution saturated in atmospheric CO2 and containing l O- 3 M total phosphates.

salt, it is known that uranium forms insoluble calcium uranates C a U O 4 and Ca3UO6 (Grenthe et al., 1992), which are mixed calcium and dioxouranium oxides that could probably add to the uranium removal at high pH values.

2.4. Long-term stability of uranium removal by HAP A series of experiments was conducted in order to evaluate the long-term stability of HAP loaded with uranium (VI) under various conditions: HAP (1.5 g/l) was added to a solution of UO2(NO3)2-6H20 (1 mg U/I at pH 9). A slurry of uranium-loaded HAP (2 g/l with approx. 650 mg U/kg HAP) was stirred in neutral (pH 8.4), alkaline (pH 13.1) and acidic solutions (pH 2.5). Samples were taken after different time intervals and analysed for uranium. The pH values of the different slurries were adjusted to the desired value using a concentrated solution of NaOH or nitric acid, respectively. The results obtained from this set of experiments are displayed in Figure 3.14. They support the assumption of a uranium bulk precipitation process by HAP for the removal of uranium from aqueous solutions. In acidic solutions (pH 2.5) the uranium dissolution is negligible even after 3 weeks of contact. It has to be noted that the addition of hydroxyapatite to the acidic solution caused a considerable increase in pH (pH is buffered to 6.3 in the pulp), due to the dissolution of HAP (see Reaction (3.3)). However, uranyl may have initially gone into solution at acidic pH and then resorbed/reprecipitated as the pH was buffered to 6.3. In neutral solutions, the amount of uranium dissolved is also negligible within a period of 3 weeks. In high-alkaline solutions (pH 13) an average of 31.6% of precipitated uranium is readily dissolved (Fig. 3.14) and remains in solution. The amount dissolved (about 265 txg/l) is well above the proposed drinking water limit (Merkel and Sperling, 1998). The behaviour of uranium dissolution in highly alkaline solutions and in neutral solutions is in good agreement with the solubility calculations for uranyl phosphate and autunite presented in Figure 3.12. In the pH region between 4 and 9, the solubility of

66


(a) "O >

"5 oo ._~ "O E ._ t-

L

D

(b)

,40

35

30 25 20 15 10 5

O . I - t.n~ 0

9

9

9

_k

350 300 v 250

. . . . . . .

....

O

9

[] [~] ,

200

[]

~

13

,

400 600 Time (h)

,~1

pH 2.5

pH 7

9 pH 13 ~0

800

100O

~ 200 E .-= 100 = 5O O

=,

9

9

_A

--i ........... ~

0

200

;1 [] 400 600 Time (h)

.[]

pH pH72.5 9 pH 13 - - - limit 800

1000

Figure 3.14. Remobilisation experiments with uranium-loaded HAP in acidic, alkaline and neutral solutions. (a) Fraction uranium dissolved and (b) corresponding uranium concentration in Ixg/l, proposed limit for drinking water (20 lxg/l) indicated as dashed line. The pH values given in the legend are initial values of the batch solutions, final pH values measured at the end of the experiments were 6.3, 7.8 and 13.1, respectively.

uranium is very low and therefore the dissolution of these phases does not occur and no uranium is transferred to the solution. As the solution pH moves to higher values, the solubility of uranium increases sharply, and therefore uranium is transferred to the solution. 2.5. Mechanism of U(VI) removal by HAP The absence of any kind of dependency of uranium removal on the solution pH, the retention time and the solid/liquid ratio (Fig. 3.7) suggests that the mechanism of uranium removal could be a typical bulk precipitation process at pH values up to 9. An adsorption mechanism in this pH region is unlikely because in that case an amount of the loaded uranium should have been desorbed regardless of the solution acidity. However, at pH values above 9 the proposed precipitation mechanism is not well supported. In this pH region, uranium occurs in the form of negatively charged aqueous species (Fig. 3.11) that could be adsorbed onto the CaOH ~ surface species (Fig. 3.9), although the overall HAP surface charge is negative ( < - 20 mV, see Fig. 3.10) and such a process is therefore energetically not favoured. For a solution pH below 3.5, uranium occurs in the form of neutral and positively charged complex uranium-phosphate species (Fig. 3.11). In the same pH region, neutral phosphate sites are present on the HAP surface along with positively charged calcium surface complexes (Fig. 3.9), resulting in the development of a strong positive charge on the HAP surface (Fig. 3.10). Due to the strong repulsive forces that arise between the positively charged species and the positively charged HAP surface, adsorption of these species is rather unlikely to happen. Adsorption is also difficult but not impossible for the neutral species, since their small mobility due to their large size presents an obstacle to their approach at the HAP surface. In the pH region between 3.5 and 5.5 uranium arises in the solution mainly in the form of the uranyl ion (Fig. 3.11) while the surface charge of HAP remains highly positive. As a result, repulsive forces arise between the uranyl ion and the HAP surface, and adsorption is unfeasible. However, some uranyl ions could reach the HAP surface due to their small ionic radius and their high mobility consuming some of the PO- sites.

Materials and processes for uranium removal from contaminated water (a) ~

(b) o~" 30

20

v

>15

..................................

E .~ 1 0

.......

O

-* .........................

O

--~ 2 5

.............

S 20

............................

E15

..................................

t_

E1O

._= 5 r D

67

o

2

4

6 pH

8

lo

12

-~ . . . . . . . . . . . . . . . . . . . -*- . . . .

..................................

E

5

5

o

..................................

o

+ AC

concentration

10

12

(g/I)

Figure 3.15. (a) Effect of pH on the uranium attenuation by activated carbon (solid/liquid ratio 2 g/l, contact time 2 h, ambient temperature). (b) Effect of the solid/liquid ratio on the uranium uptake by activated carbon (pH l 1, contact time 2 h, ambient temperature).

It is therefore difficult to reach a definitive conclusion about the removal processes for uranium with HAP because both adsorption and bulk precipitation mechanisms could contribute to the uranium removal. 3. Activated carbon Activated carbon (AC) is widely used as a sorbent in environmental applications due to its high adsorption capacity towards gaseous or dissolved compounds (Sontheimer et al., 1988). There are also a variety of applications that utilise activated carbon for groundwater remediation using PRBs; however, the target species are typically organic pollutants (Schad and Grathwohl, 1998; Tiehm et al., 2000). Adsorption of inorganic species is a process that is highly dependent on the solution pH and on the availability of adsorption sites on the reactive material. Batch experiments showed only low removal of uranium by AC (Fig. 3.15). This result is in agreement with the findings of Bostick et al. (2000) for Mesorb-3, a charcoal impregnated with approx. 10% of sulfur. The distribution coefficient Kd measured by Bostick et al. (2000) showed values around 100 l/kg, almost independent from nitrate concentration. Because of the low uranium sorption, activated carbon was not studied further. 4. Hydrated lime Hydrated lime (Ca(OH)2) is a cheap reagent which can be used in groundwater remediation. Some experience already exists in the application of hydrated lime for the remediation of acid mine drainage (Gallagher, 1998). However, it has been shown in experimental studies that hydrated lime can also be used for uranium removal from contaminated water (Cs6vfiri et al., 2002). The removal mechanism is mainly based on the increase in pH by the addition of hydrated lime. Geochemical calculations have been performed using the programme PHREEQC (Parkhurst and Appelo, 1999) to model the process. Appendix 3A lists the input file for the programme. Hydrated lime was inserted in the calculation as an infinite solid. Precipitation of oversaturated solids (such as schoepite UO2(OH)2 , beta-schoepite, and calcite CaCO3) was allowed. The calculation showed that the pH of the solution increased to 12.4 (limited

68

D. Panias, A. Xenidis, A. Krestou Table 3.2. Uranium concentration in solution calculated using the geochemical programme PHREEQC.

Uranium species

Concentration (mol/l)

UO2OH+

2.29 x 10-13 6.40 • 10 -14 1.27 • 10 -16 3.23 X 10-17 6.91 x 10-18 2.44 • 10 -20 4.63 X 10 -21

(UO2)3(OH) +

U02(C03) 2U02C03 U02(C03) 4U022+ (UO2)z(OH)2+

only by the maximum solubility of hydrated lime). The concentration of uranium was reduced to 4.2 x l 0-13 mol/l (Table 3.2).

5. Elemental iron

The application of elemental iron (Fe o) for the removal of uranium was first reported by Cantrell et al. (1995). Batch experiments presented here indicate that elemental iron can efficiently remove uranium (VI) from aqueous solutions, especially under neutral and alkaline conditions. Under acidic conditions the uranium removal by elemental iron increased steadily from approx. 82% at pH 3 to about 97% at pH 5. A further increase in the pH did not alter the uranium removal significantly, which remained almost constant at approx. 98% (Fig. 3.16a). The results also show that the reaction is quite fast: 2 h was sufficient for an almost complete removal (Fig. 3.16b). The rate of uranium attenuation was almost independent of the available surface of the reactive material (Fig. 3.16c). The presence of 400 mg/l of carbonate resulted in a 35% reduction of the uranium (VI) removal, while the presence of sulfate did not influence uranium (VI) removal (Fig. 3.16d). A series of additional batch experiments was conducted in order to evaluate the remobilisation of uranium removed by elemental iron under the various conditions described above. For the remobilisation experiments, uranium-loaded iron was prepared by adding 1.5 g/l elemental iron to a solution of UO2(NO3)2.6H20 (l mg U/I at pH 7). The product (2 g/l with approx. 650 mg U/kg iron) was then stirred in neutral/slightly acidic (pH 5.5) and alkaline (pH 13.4) water. The pH values of the different slurries were adjusted to the desired values using a concentrated solution of NaOH or nitric acid. Samples were taken after different time intervals and analysed for their uranium concentration. The results of these experiments are displayed in Figure 3.17. Around 5 - 7 % of the uranium load could be remobilised by the strongly alkaline solution from the uraniumloaded samples under the specific experimental conditions leading to a uranium concentration in the solution of around 80 I,g/l. Under neutral/slightly acidic pH conditions no significant amounts of uranium could be remobilised after an initial desorption/re-sorption step.

69


_~~

; ~ :

(b) Effect of contact time (pH 7) ~100

(a) Effect of pH variation (contact time 2 h)

,oo

-~

80

. . . . . . . . . . . . . . . . . .

O

60

.................................

E 40 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

,804o 6o

"E

E

E

t~

~

20

.................................

20

:::9

o

! i!i i!i i i i i i! !

0. 0

.

. 400

pH (c) Effect of the solid/liquid ratio (pH 7, contact time 1 h)

~ioo

-~-- . . . .

-~-

>

O

95

9O E ._..a 85 t'-80

'- 40

._~:~ 1

1600

iiiii~iiiiii

~o~

0

1200

iiiii~iiiiii

~o " 6080

.................................

E

D

800 Time (h)

(d) Effect of the presence of carbonates and sulfates (pH 7, contact time 2h)

--_ .--

I,- ....

.

2 3 4 Solid/liquid ratio (g/I)

Pure U(VI) Carbonate Sulfate Solution (400 rag/l) (200 rag/l)

5

Figure 3.16. Effect ofvarious parameters on the U(VI) removal from aqueous solution by elemental iron. All experiments were conducted at ambient temperature and a solid/liquid ratio of 2 g/l (except from c, where the solid/liquid ratio was varied), with a solution containing 1000 Ixg/l of uranium (VI).

For a discussion of the u r a n i u m attenuation m e c h a n i s m s by e l e m e n t a l iron, the iron corrosion processes which are inherent to elemental iron in aqueous e n v i r o n m e n t have to be considered. Iron corrosion depends largely on the conditions prevailing in the aqueous m e d i u m and is initiated as shown in the following reactions: Fe o + 1 0 2 + H 2 0 ~ Fe e+ + 2 0 H -

(in oxic environments)

(3.6)

Fe ~ + 2 H 2 0 ~ Fe 2+ + H2 + 2 O H -

(in anoxic environments)

(3.7)

(b)

(a)~l 0 "O

r

8

..................................

.a o 6 E 4

9

D

0

9

.8ot.o: 6o .............. -

9

O

E 2~ -,._9

9

~100

.

.

.

.

o pH

_~ ................ 0

200

400

. 600

Time (h)

5.5_ I

pH13.41

800

1000

...."....:..... I

E 40-;-- . . . . . . . . . . . . . . . . . . . . . . . . . . . .

o pa 5 5 - I pH 13.4_1

._.a = 20 ~

|

Z)

O-P--o

0

~,

200

9

,

400

__ ,

600

limit ,

800

l I

1000

Time (h)

Figure 3.17. Amount of uranium remobilised in alkaline and neutral solutions from uranium-loaded elemental iron: (a) % uranium remobilised and (b) corresponding uranium concentration in Ixg/l (proposed limit for drinking water of 20 txg/l shown by dashed line).

70


The above reactions are always accompanied by the formation of precipitates (Blowes et al., 2000; Roh et al., 2000; Qiu et al., 2000; Furukawa et al., 2002; Morrison et al., 2002) such as amorphous iron oxides and oxyhydroxides, goethite, aragonite, calcite, siderite, etc. as described in the following reactions: 2Fe 2+ + 102 -~- 3H20 *-" 2FeOOH(solid) + 4H +

(3.8)

2Fe2+ + 1 02 + 2H20 ,__,Fe203(solid) + 4H +

(3.9)

Fe 2+ + CO 2- ,.., FeCO3(solid)

(3.10)

The potential of such precipitates to negatively affect the overall function of an Fe ~ barrier has already been discussed in Chapter 1 where some results from field installations are also reviewed. On the other hand, the precipitated minerals offer additional surface areas within the barrier and can thus contribute to immobilisation of certain contaminants through sorption or co-precipitation. For the removal of uranium, three possible mechanisms can facilitate uranium (VI) attenuation: reduction of uranium (VI) to uranium (IV), precipitation of U(VI) as uranium hydroxide, and uranyl adsorption onto iron oxide surfaces. 5.1. Reduction of uranium (VI) to uranium (IV) The reduction mechanism implies the reduction of U(VI) to U(IV) on the surface of elemental iron and precipitation of the sparingly soluble uraninite, as indicated in the following reaction (Abdelouas et al., 1999; Blowes et al., 2000; Roh et al., 2000; Qiu et al., 2001): Fe ~ + UO 2+ ~ Fe 2+ +

UO2(solid)

(3.11)

Blowes et al. (2000) have shown that strongly reducing conditions must be attained to enable the above reactions to proceed. Simon and Meggyes (2000) have reported that the above reaction could be affected by the presence of CO2, which is a typical constituent in groundwater. Additionally, Qiu et al. (2001) reported that even if the above reaction takes place on the surface of elemental iron, the newly formed solid uraninite (U02) acts as a nucleation site for the initiation of continuous precipitation of the amorphous U(VI) hydroxide (schoepite). 5.2. Precipitation of U(VI) as an amorphous uranium hydroxide The continuous corrosion of elemental iron in the solution (Reactions (3.6) and (3.7)) increases the local pH around the iron particles. Therefore, as the local pH in the vicinity of the Fe~ surface is higher than the pH in the bulk of the solution, precipitation of amorphous uranium (VI) hydroxide (schoepite) can take place close to the iron surface (Qiu et al., 2000, 2001). Therefore, uranium can be removed from the solution also through a nonreductive precipitation mechanism. This mechanism is affected by the presence of C02 in the solution. Carbonates form very stable soluble complexes with uranyl ions, especially in the alkaline region, which do


71

not form strong bonds with the iron oxides that are present on the iron surface. Thus, carbonates can inhibit the growth of amorphous uranyl hydroxides on the Fe ~ surface. 5.3. Uranyl adsorption onto iron oxides

The corrosion of iron results in the formation of several amorphous iron oxides and oxyhydroxides that could act as adsorbents (Gu et al., 1998) for the uranyl ions. To conclude, there are three possible pathways that are responsible for the uranium (VI) attenuation from groundwater using elemental iron. Although the reductive mechanism is feasible, it needs strongly reducing conditions to proceed. On the other hand, the precipitation of uranyl hydroxides along with the sorption of the uranyl ions on the iron oxyhydroxides seems to be more feasible under the conditions prevailing in groundwater. Especially, the U(VI) removal through sorption reactions onto the iron oxides seems to be the most important one. The experimental results show that the U(VI) removal equilibrates rapidly and reaches equilibrium within a few hours (Fig. 3.16b), which might favour an adsorption mechanism rather than precipitation. 6. Iron oxides

Uranium can be removed from aqueous solution by adsorption on reactive surfaces. Morrison and Spangler (1992) evaluated a range of uranium and molybdenum adsorption tests using a variety of materials. Good removal results have been obtained for lime, hematite, peat, ferric oxyhydroxides, phosphate and TiO2, while clays exhibited only low sorption potential. Precipitation and surface adsorption are processes which can occur simultaneously in a chemical barrier. Sorption of uranium from groundwater was studied in a series of papers (Morrison and Spangler, 1992, 1993; Morrison et al., 1995). Surface site complexation can be described using different models with or without electrostatic influence on charged surfaces (Allison et al., 1991). With S - O H as a notation for a surface site the adsorption reaction of uranium onto ferric oxyhydroxide as a sorbent can be written as follows: S - O H 4- UO22+ *-+ (SO)UO + 4- H +

(3.12)

Sorption, rather than precipitation, depends strongly on pH. Therefore, it might happen that mixed contaminants cannot be remediated with sorption processes. For example, uranium cannot be removed simultaneously with molybdenum because the latter is mobile at pH values above 8 while uranium exhibits low mobility at this pH in the absence of carbonate (Morrison and Spangler, 1993). Both sorption onto surfaces and abiotic reduction are possible removal mechanisms in iron-beating walls. The amount and efficiency of uranium removal by Fe~ barriers through these distinct processes have been studied in several articles (Fiedor et al., 1998; Gu et al., 1998). It was found that Fe ~ filings are much more efficient than pure adsorbents. A recent technical report (Office of Radiation and Indoor Air, 1999) reviews the properties of the aqueous solution and sorbent that are most important in controlling adsorption/retardation behaviour of uranium and other environmentally relevant elements. The distribution coefficient Kd of uranium at different pH values in various soils was

72

D. Panias, A. Xenidis, A. Krestou 7

5

9

.

o

9

"* . [ ' . .

olL.'

,-

.**

9

-1 -2

2

&

&

;

1'o

11

pH Figure 3.18. Distributioncoefficient Ko of uranium as a function of pH (source: Office of Radiation and

Indoor Air, 1999). compiled from literature. The data are displayed in Figure 3.18. The data exhibit a large amount of scatter but show a trend as a function of pH. High Ka values were derived from adsorption experiments with ferric oxyhydroxide and kaolinite, low values from those with quartz which has low adsorptive properties. The pH dependency arises from surface charge properties of the mineral compounds and from the complex aqueous speciation of uranium (VI). The dissolved carbonate concentration also has a significant influence due to the formation of stable uranyl-carbonato complexes. Additional data are given by Bostick et al. (2000) who measured Ka values for the sorption of uranium to Apatite II (a reactive form of HAP) of nearly 105 l/kg at neutral pH whereas values around l OO l/kg were estimated for activated carbon.

D. Summary and conclusions The experimental results indicated that adsorption is the principal removal process for uranium (VI) by zeolites. Although the studied zeolitic tuff removed up to 90% of the initial uranium concentration from aqueous solution in batch experiments, only poor removal of uranium (VI) was observed in the presence of carbonates. The zeolite uranium removal capacity is only a small fraction of the CEC because uranium cannot access the interior of the zeolite due to its size. Moreover, the efficiency of zeolites as reactive materials depended markedly on the solution pH. Hydroxyapatite is a promising reactive material regarding its use in the remediation of uranium-contaminated groundwater. The attenuation process by HAP could be attributed to bulk precipitation, especially for solution pH up to about 9, while sorption onto HAP surface occurs at higher pH values. The removal products are very stable in the neutral pH range, but unstable at high alkaline values where a large amount of uranium can be remobilised. Hydroxyapatite showed a high (about 95%) removal efficiency in the pH range of 3 - 1 l, and the observed efficiency was independent of the presence of carbonates and sulfates (the most common anions in groundwater). Hydroxyapatite exhibited


73

a removal capacity of at least 20 mg U/g HAP and an ability to control the groundwater pH especially in the range of 5 . 5 - 9 . However, hydroxyapatite is gradually consumed due to its solubility in water. Although hydrated lime appears to be an effective material for uranium attenuation, as documented by a high uranium removal from aqueous solution in batch experiments, the application of the material may cause problems during longer operational periods due to dissolution of the material and major increase in pH, which may be harmful to the environment. Elemental iron removes uranium (VI) very effectively from aqueous solutions. However, the process is negatively affected by the presence of dissolved carbonates. The results suggest that the mechanism is a combination of adsorption and precipitation. A small amount of the removed uranium (about 2% under the specific experimental conditions described above) can be remobilised in water at pH 5.5.

References Abdelouas, A., Lutze, W., Nuttal, E., Gong, W., 1999. Rrduction de U(VI) par le fer m&allique: application ?~la drpollution des eaux. Grosciences de Surface, 315-319. Allison, J.D., Brown, D.S., Novo-Gradac, K.J., 1991. Minteqa2/Prodefa2, A Geochemical Assessment Model for Environmental Systems. US Environmental Agency, Database of computer programme, Version 3.0, Athens, GA, USA. Blowes, D.W., Ptacek, C.J., Benner, S.G., McRae, C.W.T., Bennet, T.A., Puls, R.W., 2000. Treatment of inorganic contaminants using permeable reactive barriers. J. Contam. Hydrol. 45, 123-137. Bostick, W.D., Stevenson, R.J., Jarabek, R.J., Conca, J.L., 2000. Use of apatite and bone char for the removal of soluble radionuclides in authentic and simulated DoE groundwater. Adv. Environ. Res. 3, 488-498. Brown, D., Potter, P.E., Wedemeyer, H., 1981. Uranium, supplement volume C14. In: Keim, R. (Ed.), Gmelin Handbook of Inorganic Chemistry. Springer, Berlin. Cantrell, K.J., Kaplan, D.I., Wietsma, T.W., 1995. Zero-valent iron for the in-situ remediation of selected metals in groundwater. J. Hazard. Mater. 42, 201-212. Csrvfiri, M., Csics~ik, J., Frlding, G., 2002. Investigation into calcium-oxide based reactive barriers to attenuate uranium migration. In: Simon, F.G., Meggyes, T., McDonald, C. (Eds), Advanced Groundwater Remediation - Active and Passive Technologies. Thomas Telford, London. Fetter, C.W., 1993. Contaminant Hydrogeology. Prentice-Hall, Upper Saddle River, p. 500. Fiedor, J.N., Bostick, W.D., Jarabek, R.J., Farrell, J., 1998. Understanding the mechanism of uranium removal from groundwater by zero-valent iron using X-ray photoelectron spectroscopy. Environ. Sci. Technol. 32, 1466-1473. Furukawa, Y., Kim, J.-W., Watkins, J., Wilkin, R.T., 2002. Formation of ferrihydrite and associated iron corrosion products in permeable reactive barriers of zero-valent iron. Environ. Sci. Technol. 36, 5469-5475. Gallagher, D., 1998. Groundwater Pollution Primer, Civil Engineering Department, VirginiaTech, Internet Report, CE 4594. Gottardi, G., Galli, E., 1985. Natural Zeolites. Springer, Berlin, 409 p. Grenthe, I., Fuger, J., Konings, R.J.M., Lemire, R.J., Muller, A.B., Nguyen-Trung, C., Wanner, H. 1992. Chemical thermodynamics of uranium. In: Wanner, H., Forest, I. (Eds), Chemical Thermodynamics, VoI. 1, North-Holland, Amsterdam. Gu, B., Liang, L., Dickey, M.J., Yin, X., Dai, S., 1998. Reductive precipitation of uranium (VI) by zerovalent iron. Environ. Sci. Technol. 32, 3366-3373. Gustafsson, J.P., 2003. Visual Minteq, Version 2.22, Software, http://www.lwr.kth.se/English/ OurSoftware/vminteq/.

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Krestou, A., Panias, D., Xenidis, A., Paspaliaris, I., 2002. Uranium removal from aqueous solutions using natural zeolite. In: Proceedings of the Seventh International Symposium on Environmental Issues an Waste Management in Energy and Mineral Production (SWEMP 2002), 7-10 October 2002, Cagliari, Sardinia, Italy, pp. 795-802. Krestou, A., Xenidis, A., Panias, D., 2003. Mechanism of aqueous uranium (VI) uptake by natural zeolitic tuff. Miner. Eng. 16, 1363-1370. Krestou, A., Xenidis, A., Panias, D., 2004. Mechanism of aqueous uranium(VI) uptake by hydroxyapatite. Miner. Eng. 17, 373-381. Langmuir, D., 1978. Uranium solution-mineral equilibra at low temperatures with applications to sedimentary ore deposits. Geochim. Cosmochim. Acta 42, 547-569. Lepperd, D., 1990. Heavy metal sorption with clinoptilolite zeolite: alternatives for treating contaminated soil and water. J. Mining Eng. June, 604-608. Leyva, A., Marrero, J., Smichowski, P., Cicerone, D., 2001. Sorption of antimony onto hydroxyapatite. Environ. Sci. Technol. 35, 3669-3675. Loizidou, M., Townsend, R.P., 1987. Ion-exchange properties of natural clinoptilolite ferrierite and mordenite: Part II. Lead-sodium and lead-ammonium equilibria. Zeolites 7, 153-159. Merkel, B., Sperling, B., 1998. Hydrogeochemische stoffsysteme II. Schriftenreihe des Deutschen Verbandes ftir Wasserwirtschaft und Kulturbau e.V. (DVWK), VoI. 117, Deutscher Verband ftir Wasserwirtschaft und Kulturbau e.V. (DVWK), Bonn. Morrison, S.J., Spangler, R.R., 1992. Extraction of uranium and molybdenum from aqueous solutions: a survey of industrial materials for use in chemical barriers for uranium mill tailings. Environ. Sci. Technol. 26, 1922-1931. Morrison, S.J., Spangler, R.R., 1993. Chemical barriers for controlling groundwater contamination. Environ. Prog. 12, 175. Morrison, S.J., Spangler, R.R., Tripathi, V.S., 1995. Adsorption of uranium (VI) on amorphous ferric oxyhydroxide at high concentrations of dissolved carbon (IV) and sulfur (VI). J. Contam. Hydrol. 17, 333-346. Morrison, S.J., Metzler, D.R., Dwyer, B.P., 2002. Removal of As, Mn, Mo, Se, U, V and Zn from groundwater by zero-valent iron in a passive treatment cell: reaction process modelling. J. Contam. Hydrol. 56, 99-116. Naftz, D.L., Morrison, S.J., Davis, J.A., Fuller, C.C., 2002. Handbook of Groundwater Remediation using Permeable Reactive Barriers. Elsevier, Amsterdam, 539 p. Office of Radiation and Indoor Air and Office of Environmental Restoration, 1999. Review of geochemistry and available Ko values for cadmium, cesium, chromium, lead, plutonium, radon, strontium, thorium, tritium and uranium. US Environmental Protection Agency and US Department of Energy, Understanding Variation in Partition Coefficient, Ka Values, VoI. II, EPA 402-R-99-OO4B, Washington, DC, USA. Onal, G., Atak, S., Gtiney, A., ~elik, M.S., Ytice, A.E., 2001. New developments in mineral processing. In: Proceedings of the Ninth Balkan Mineral Processing Congress, Istanbul/Turkiye, 11-13, September 2001. Parkhurst, D.L., Appelo, C.A.J., 1999. User guide to Phreeqc (version 2) A computer program for speciation, batch-reaction, one-dimensional transport and inverse geochemical calculations. US Geological Survey, Water-Resources Investigations Report, 99-4529, Denver. Qiu, S.R., Lai, H.-F., M, J., Hunt, M.L., Amrhein, C., Giancarlo, L.C., Flynn, G.W., Yarmoff, J.A., 2000. Removal of contaminants from aqueous solution by reaction with iron surfaces. Langmuir 16, 2230-2236. Qiu, S.R., Amrhein, C., Hunt, M.L., Pfeffer, R., Yakshinskiy, B., Zhang, L., Madey, T.E., Yarmoff, J.A., 2001. Characterization of uranium oxide thin films grown from solution onto Fe surfaces. Appl. Surf. Sci. 181, 211 - 224. Regnier, P., Lasaga, C.A., Berner, R.A., Han, O.H., Zilm, K.W., 1994. Mechanism of CO~- substitution in carbonate-fluorapatite: evidence from PTIR spectroscopy, 13C NMR, and quantum mechanical calculations. Am. Miner. 79, 809-818. Roh, Y., Lee, S.Y., Elless, M.P., 2000. Characterization of corrosion products in the permeable reactive barriers. Environ. Geol. 40, 184-194. Schad, H., Grathwohl, P., 1998. Funnel-and-gate systems for in situ treatment of contaminated groundwater at former manufactured gas plant sites. In: NATO/CCMS Special Session on Treatment Walls and Permeable Reactive Barriers, Vienna, Austria, EPA 542-R-98-003, pp. 56-65.


75

Simon, F.G., Meggyes, T., 2000. Removal of organic and inorganic pollutants from groundwater using permeable reactive barriers - part 1: treatment processes for pollutants. Land Contam. Reclamation 8, 103-116. Sontheimer, H., Crittenden, J.C., Summers, R.S., 1988. Activated Carbon for Water Treatment, 2nd edn. DVGW-Forschungsstelle Karlsruhe, 722 p. Tiehm, A., Schulze, S., B6ckle, K., Miiller, A., Lorbeer, H., Werner, P., 2000. Elimination of chloroorganics in a reactive wall system by biodegradation on activated carbon. In: Proceedings of ConSoil 2000, September 18-22, Leipzig, Germany, Thomas Telford Publication, London, pp. 924-931. Wu, L., Forsling, W., Schindler, P.W., 1991. Surface complexation of calcium minerals in aqueous solution, 1. Surface protonation of fluorapatite-water interfaces. J. Colloid Interf. Sci. 147, 178-185.


Long-term Performance of Permeable Reactive Barriers K.E. Roehl, T. Meggyes,F.-G. Simon, D.I. Stewart, editors 9 2005 Elsevier B.V. All rights reserved.

77

Chapter 4 Behaviour of uranium in elemental iron and hydroxyapatite reactive barriers: column experiments Franz-Georg Simon and Vera Biermann

A. Introduction Uranium is present in groundwater and surface waters as a result of natural processes and anthropogenic activities such as uranium mining and milling. The uranium concentration in surface waters arising from natural processes averages at 0.2 t.~g/l, with local peak values of 200 ~g/l (Fauth et al., 1985). Uranium concentration in groundwater is much higher where human activities have resulted in the contact of water with uranium deposits. Concentrations in the order of several mg/l have been observed (Morrison and Spangler, 1992; Naftz et al., 1996; Cs6v~ri et al., 2002). In such cases groundwater remediation measures are necessary to reduce the uranium concentration in the groundwater to the proposed limit for uranium in drinking water of 20 p~g/l (Merkel and Spefling, 1998). In practice, water treatment facilities use ion exchangers which selectively remove uranium from the water (Merkel and Sperling, 1998; Simon and Meggyes, 2000). A variety of reactive materials have been proposed and tested for uranium removal from groundwater. The application of elemental iron was first investigated by Cantrell et al. (1995). Good results have also been obtained for uranium adsorption onto phosphate minerals, lime, hematite, peat, TiO2, and ferric oxyhydroxide, while clays exhibited only low sorpfion potential (Morrison and Spangler, 1992). For the removal of uranium from groundwater elemental iron (Fe o, often termed as zero-valent iron or ZVI in this context) and the phosphate mineral hydroxyapatite (HAP) are the favourite candidates for use in permeable reactive barriers (PRBs). The investigations reported in this chapter were therefore focussed on these two substances. At Fry Canyon in Utah, USA, a field demonstration of in situ chemical barriers has been in operation to control uranium contamination in groundwater since 1997 (Naftz et al., 1999, 1996). Here elemental iron, amorphous iron oxide and bone char phosphate containing hydroxyapatite have been used as active components to remove uranium from the groundwater. Both iron and hydroxyapatite showed good retention abilities over several years of operation with more than 99% uranium being removed from the groundwater. However, removal efficiency of amorphous iron oxide decreased within the first year (Naftz et al., 1999). Uranium removal from groundwater can be performed using active methods (pumpand-treat) or passive in situ methods (PRBs) (Simon et al., 2002). The processes applied

78

F.-G. Simon, V. Biermann

for removal are the same in both technologies and imply precipitation, ion exchange, sorption and chemical reduction. Uranium mainly occurs in the oxidation states + 4 and + 6. Hexavalent uranium, i.e. the uranyl ion UO~ +, is more mobile in the aqueous environment than U(IV) compounds. Good removal results have been obtained by reduction of U(VI) to U(IV) using elemental iron Fe + UO22+(aq) ~ Fe 2+ +

UO2(s)

(4.1)

The ability of elemental iron to reduce the uranyl ion UO 2+ was first reported by Cantrell et al. (1995). The mechanistic aspects of uranium removal from groundwater using Fe~ are still being discussed. Possible reaction paths are reductive precipitation or adsorption onto the corrosion products of Fe ~ (Gu et al., 1998). Reduction of U022+ by elemental iron is possible over a wide range of pH as illustrated in Figure 4.1. The solubility of uraninite (U02) is around lO -s mol/l in a pH range between 4 and 14. Below pH = 4 uraninite becomes soluble. The behaviour of uraninite has been studied intensely because of its application as a nuclear fuel and since it is a main constituent of the uranium mineral pitchblende (Bickel et al., 1996). Measurements on the dissolution of spent fuel in deionised water under non-oxidising conditions yielded solubilities in the range of 10-9-10 -5 mol/l. Solubility is enhanced under oxidising conditions where UO2 can be transformed into the uranyl ion and where complexation reactions may occur. It is therefore crucial that Eh is below the boundary of the U(IV)/ U(VI) curve in the U(IV) stability field in Figure 4.1 if reductive precipitation should be the predominant removal mechanism. With C02 present, the boundary line is shifted towards lower Eh at high pH. This behaviour is also displayed in Figure 4.1. It is therefore more difficult to transform uranium into the oxidation state + 4 under alkaline conditions in the presence of C02 (O.Ol bar C02 is a typical value for groundwater conditions). A review on complex uranium solution equilibria is given by Langmuir (1978). The technology of using elemental iron for groundwater remediation was developed and patented by the University of Waterloo in Canada (Gillham, 1996). Iron was proven to be particularly effective for the treatment of groundwater polluted by chlorinated hydrocarbons. Sivavec and Homey (1995) tested 25 different iron samples from various 1000 800600>

g U.J

4002000-200 -400 -600

0

Fe/Fe 2+

P/-12;

~

C

U(lV) .................... 2 i "..............................

0

I

I

I

I

2

4

6

8

10

pH Figure 4.1. Eh/pH - diagram showing the boundary line between U(IV) and U(VI) without and in the presence of 0.01 bar C02 (Simon and Meggyes, 2000).

Behaviour of uranium in elemental iron and hydroxyapatite reactive barriers

79

sources for their ability to facilitate degradation of chlorinated compounds. They and other scientists (Matheson and Tratnyek, 1994; Agrawal and Tratnyek, 1996) found that the effectiveness of these iron samples mainly depends on the maximum available specific surface area (m2/g). In an aqueous environment the oxidation ofFe ~ to Fe 2+ (Reaction (4.2)) is coupled with the reduction of water (Reaction (4.3)), and thus causes an increase in pH which might favour precipitation of minerals depending on the constituents of the water and the residence time in the PRB Fe ---, Fe 2+ + 2e-

(4.2)

2H20 + 2e- ---, H2 + 2 0 H -

(4.3)

If precipitation of minerals occurs, the effective pore volume over a cross-section of the PRB is reduced. In this case the hydraulic conductivity decreases, resulting in a reduction of the active Fe ~ surface and its reactivity. High dissolved oxygen content in the water also promotes an increase in pH making the situation even worse: 02 + 2H20 + 4e- ~ 4 0 H -

(4.4)

Previous investigations (Mackenzie et al., 1997; 1999) showed that reaction with oxygen takes places at the entrance of the reactive zone. Iron oxides and oxyhydrates are precipitated which results in a decline of permeability. Not restricted to the entrance zone but spread through the barrier is the release of H2 due to the oxidation ofFe ~ in the absence of oxygen, filling the pores and contributing to a decrease in hydraulic permeability. The form of the particles of the reactive material is also of importance to the reactions. Steel fibres from shredded tyres have similar chemical properties as iron particles. However, the specific surface area of steel fibres is much smaller. The expected efficiency is therefore lower than that of elemental iron particles (granules). However, steel fibres/sand mixtures have a larger pore volume leading to a better permeability, and the effect of precipitation and aggregation of particles due to corrosion is negligible here. Steel fibres are waste materials and are therefore very cheap. The application of apatite minerals (Cas(PO4)3X, X = halide, hydroxyl) has proven effective in immobilising many transition and heavy metals and radionuclides through the formation of secondary phosphate precipitates that remain stable over a wide range of geochemical conditions (Seaman et al., 2001). The structure of hydroxyapatite can be illustrated as displayed in Figure 4.2. The uranium concentration in a solution can be reduced without changing the oxidation state by forming uranium phosphates of low solubility, e.g. precipitating uranyl ions with phosphate forming (UO2)3(PO4) 2 (log K s p - - - 49.09; Brown et al., 1981) or with HAP or bone char (HAP with a small amount of carbon) forming autunite Ca(UO2)z(PO4)2 (log K s p = - 4 7 . 2 8 ; Brown et al., 1981) or chernikovite H2(UO2)2(PO4) 2 (log K s p - - - 4 5 . 4 8 ; Grenthe et al., 1992). The mechanism of the interaction of uranium with HAP is not completely understood. Jeanjean et al. (1995) proposed a dissolutionprecipitation mechanism. If autunite or chernikovite is the precipitation product the reaction sequence would be: Cas(PO4)3OH--~ 5Ca 2+ + 3P03- + OH-

(4.5)

80


Figure 4.2. Structure of hydroxyapatite (Remy, 1959). Water molecules are not displayed.

H + + OH- ~ H2O

(4.6)

2U022+ + Ca 2+ q- 2P043- ---"Ca(UO2)2(PO4)2

(4.7)

2U0 2+ + 2H + + 2P034-

(4.8)

-- H2(UO2)2(PO4)2

Ion exchange processes with HAP (see Reaction (4.9)) or surface sorption on HAP at two possible surface groups (Reactions (4.1 O) and (4.11)) can be described as follows (Wu et al., 1991; Leyva et al., 2001): ~ C a 2+ + UO~ + ~ ~UO~ + + Ca 2+

(4.9)

~ O H + UO 2+ - - ~ O - U O ~ - + H +

(4.10)

~ - 0 3 P - O H + + UO 2+ ~ ~ O 3 P - O - U O 2+ + H +

(4.11)

In the work of Fuller et al. (2OO2a) autunite and chernikovite have been identified as solid phases when uranyl ions were added to a saturated solution of hydroxyapatite. Evidence of U(VI) adsorption to hydroxyapatite surfaces as an inner-sphere complex was found for certain concentration ratios. Arey et al. (1999) postulated the formation of secondary phosphate phases in their work on uranium immobilisation by HAP in sediments but could not identify any controlling phase such as autunite. In two articles on the application of Apatite II to the remediation of metal-contaminated soils and groundwater (Conca et al., 2000; Conca and Wright, 2003) the detection of autunite phases is described. These results suggest that heavy metals and radionuclides are very efficiently concentrated on apatite with no desorption, leaching or ion exchange due to changes in pore water chemistry, pH or temperature. An investigation into the remediation of uranium-contaminated groundwater using Apatite II revealed autunite precipitation onto Apatite II surfaces as the predominant mechanism (Bostick et al., 2000). Up to now information on uranium removal from groundwater by hydroxyapatite is limited. Uranyl ions, like lead, can be precipitated to sparingly soluble phosphate compounds by hydroxyapatite. However, hydroxyapatite has a large specific surface, and therefore surface-sorption reactions may also influence the process.


81

B. Initial laboratory column test systems 1. Materials and methods

1.1. Experimental set-up Test columns to evaluate the performance of different reactive materials for uranium removal from artificial groundwater were set up as shown by the schematic diagram in Figure 4.3. A multi-channel peristaltic pump was used to pump the artificial groundwater through the columns. The flow direction was from bottom to top, with a layer of filter sand at either end to achieve better flow conditions within the reactive material. For monitoring purposes, an oxygen-sensor was integrated into the recharge tubing and pH, redox and conductivity electrodes were installed into the discharge system. The set-up also included several water sampling points and pressure gauges to observe any pressure build-up at the column head as an indicator of pore clogging and loss of permeability. Column effluents were analysed spectrophotometrically for uranium using an arsenazo III-method adapted from Savvin (1961), Singer and Matucha (1962) and Korkisch (1972). The method is based on the reduction of hexavalent uranium ions (UO 2+) to tetravalent uranium with metallic zinc prior to complexation with arsenazo III in > 4 molar hydrochloric acid. Subsequently, the sample is analysed photometrically at 666 nm wavelength. Uranium concentrations of less than 50 p,g/l cannot be analysed accurately with this method. Long-term column tests were run with five columns each 50 cm high and 6.3 cm in diameter. The height of the reactive zone was 40 cm in all columns. Two of the columns were filled with hydroxyapatite/sand mixtures, one containing 20% and the other 10% HAP by weight. The other three columns were filled with elemental iron/sand mixtures

Figure 4.3. Sketchand photograph of the test columns used in experiments on uranium removal from artificial groundwater.

82


with an iron content of 70, 50, and 30% by weight. The reactive media and the filter sand are shown in Figure 4.4. The artificial groundwater used as a feed solution had a uranium concentration of 0.49 + 0.07 mg/l. Its composition is described in detail in Section 1.3. The initial flow rate was set to 100 ml/h but the average ranged between 58 and 64 ml/h due to the wear of the peristaltic pump tubes. The experiments have been conducted as accelerated tests so that any ageing effects would be visible earlier than in natural systems. The concept of the accelerated method comprises higher groundwater flow velocities in the columns compared to natural conditions and higher concentrations of the groundwater contaminants. 1.2. Materials

Shredded cast iron (granulated grey cast iron, 0.3-1.3 mm) supplied by Gotthard Mayer, Rheinfelden, Germany, and food quality grade hydroxyapatite (Cas(PO4)3OH, 99% < 0.4 mm) supplied by Chemische Fabrik Budenheim CFB, Germany, were investigated.

1.3. Water composition

In the laboratory column experiments, artificial groundwater was used to simulate the groundwater chemistry of the case study site near Prcs, Hungary (see Chapter 9). The composition of the artificial groundwater in comparison to the original site groundwater is shown in Table 4.1. The original groundwater can be characterised as a C a - M g - H C O 3 SO4 water, with an oxygen content of up to 2 mg/l. The water chemistry was modelled using the geochemical simulation software Phreeqc (Parkhurst and Appelo, 1999) and the thermodynamic database from the computer code MinteqA2 (Allison et al., 1991), which is supplied with Phreeqc. A simulation was run with Phreeqc modelling the chemistry of the feed water to the column tests using the input data from Table 4.1. The experimental results are displayed in Appendix 4A.

Figure 4.4. Reactive materials and filter sand for mixing used in the test columns.


83

Table 4.1. Comparison of water compositions from the test site and water used in the experiments. Constituents

Unit

Column tests

Field site

U(VI) Na + K+ Ca 2+ Mg 2+ CISO2CO 2HCO3 Carbonate hardness Total hardness pH Electrical conductivity

p~g/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l ~ ~ IxS/cm

490 63.7 16.0 134 39.4 47.5 234 11.4 443 21 28 6.8 909

385 44 14 140 40 32 253 10 412 18.9 28.8 6.8 1061

Anions and cations are nearly balanced; the difference is only 3.5%. Anhydrite, aragonite, calcite and dolomite are slightly oversaturated so that precipitation of these minerals can be expected. Nearly no reduction of uranium (VI) occurs before contact with elemental iron. The predominant uranium (VI) species are U02(C03) 2- and U02(C03) 4-. This behaviour is explained in detail by Langmuir (1997).

2. Results of the laboratory column experiments In a first experiment, a test column was filled with cast iron filings with a grain size range of about 0 . 3 - 1 . 3 m m while a second column was filled with filter sand (0.25-1.0 mm). The flow rate was set to 2 ml/min and the pressure build-up at the column heads was measured. In these column tests, the pressure built up more rapidly in the iron column than in the sand column, indicating a significant loss of permeability in the iron column (Fig. 4.5). 1.0 L..

.a v

g

................. /

0.8

..........

!!!!!!!!!

-.~ 0.6 ..Q

0.4

0.2 a.

0.0

O

'

'

'

'

I

5

'

'

'

'

I

'

'

10 Time (days)

'

'

I

15

'

'

Figure 4.5. Pressure build-up in test columns of an initial experiment series.

'

'

20

84


After 18 days flow through the iron column ceased while the sand column was still operating well. It was assumed that the pore space of the iron column was clogged. Therefore, the experiment was terminated and the iron column was dismantled. A massive iron hydroxide core was discovered in the lower part of the column. This was an important experimental result, although in field PRB applications iron is always used in a mixture with sand to avoid complete clogging. All the remaining column experiments were conducted using iron/sand mixtures. Further column experiments were carried out to test the long-term performance of elemental iron and hydroxyapatite as reactive materials for uranium removal. Five columns were filled with either Fe ~ or HAP and filter sand as specified in Table 4.2. Initially, the hydraulic conductivity of the columns was about 1 0 - 5 - 1 0 -7 m/s and the flow rate was set to 1.5 ml/min which corresponds to a flow velocity of about 5 pore volumes per day. The characteristics of the feed water with an average uranium concentration of 490 +__68 l~g/l are displayed in Table 4.1. The uranium concentration, pH, redox potential, and conductivity of the column effluent, as well as the pressure at the column inlet, were monitored. No breakthrough occurred over a period of almost 2 years (01.06.2001-16.04.2003), with the average uranium retention of all columns being better than 90%. A more precise quantification of the uranium removal is not possible due to the high detection limit and variation of the analytical photometric method used. A control measurement with the ICP-MS method in August 2002 yielded effluent uranium concentrations below 0.1 Ixg/l corresponding to a uranium retention rate of at least 99.7% for both Fe ~ and HAP used as reactive materials. While both materials performed equally well with regard to uranium retention, only the HAP columns maintained their hydraulic conductivity throughout the experiment. The pressure at the inlet of the column with 70% Fe ~ started to build up after about 8 months of operation rising from an average O.1 to 0.3 bar (Fig. 4.6). This was followed by an even quicker rise up to 2.0 bar, the limit of the peristaltic pump, indicating a dramatic loss of permeability after 1 year. The same happened only i month later with the 50% Fe~ column (not displayed). Better results were achieved with the 30% Fe ~ column where the first significant sign of clogging was observed only after 22 months when the column pressure reached 0.5 bar. In comparison, the pressure at the heads of the HAP columns did not exceed 0.2 bar throughout the experiment. The two materials had different effects on the effluent pH (Fig. 4.7). While the HAP columns basically did not change the pH of the feed water from its initial value of about pH 7, the iron columns caused an increase in pH, with values of up to 9.5 in the column effluents. Table 4.2. Specifications of the reactive materials used in long-term column experiments.

Column

Reactive material (RM)

Mass ratio RM/sand

Mass of RM (g)

C2 C3 C4 C5 C6

HAP HAP Fe ~ Fe~ Fe~

20/80 10/90 70/30 50/50 30/70

306 158 1795 1160 587


85

Figure 4.6. Pressure at the column heads during the long-term experiments (01.06.2001-16.04.2003). However, this effect diminished after 2 - 3 months. Then the effluent pH of the Fe ~ columns gradually decreased to values between around 7.2 (30% Fe o) and 7.5 (70% Fe~ The findings of the long-term column experiments indicate that uranium removal with Fe ~ and HAP is based on different reaction mechanisms. The reductive precipitation mechanism postulated for Fe ~ suffers from various side reactions that cause an increase in pH, the formation of H2 and the reduction of the redox potential. As a result, the precipitation of minerals like calcite is favoured, which can result in the clogging of the pores and the impairment of the hydraulic conductivity of the reactive material. Consequently, the flow pattern inside the reactive material could change and preferential flow paths develop. The effluent pH can serve as an indicator for this process as the pH increase due to iron oxidation becomes less distinct as the residence time decreases due to preferential flow paths (Kamolpornwijit et al., 2003). In our experiments, the effluent pH of the Fe ~ columns always decreased before a pressure increase at the column head indicated clogging. This scenario cannot be directly extrapolated to a field barrier, where clogging will tend to cause by-pass flow rather than a pressure build-up. In order to determine the uranium distribution in the clogged Fe~ columns, the columns were dismantled, cut in half down the centreline and divided into segments. These segments were dissolved in concentrated hydrochloric acid, filtered, and then analysed

Figure 4.7. pH of column effluents during the long-term experiments (01.06.2001-16.04.2003).

86

F.-G. Simon, V. Biermann 40 355

..........................................

30

g 25 "

Photometry

- -~

~ 2o

-ICP-MS 9

E 10

ICP-AES

l

5 0

i

0

2000

i

i

i

i

4000

i

6000

Uranium concentration (mg U/kg Fe)

Figure 4.8. Uranium distribution in the 70% Fe~ column.

using photometry for dissolved iron and uranium. As iron interferes with the arsenazo-III method for uranium analysis, the standard addition method was used and ICP measurements were performed to validate the results. The column with 70% Fe ~ was operated for 559 days with an average flow rate of 60 ml/h (3.1 pore volumes per day), which resulted in a total input of 396 mg of uranium. The major part of that uranium was retained within the first 8 cm of the reactive material of the 70% Fe ~ column (Fig. 4.8). The uranium concentration was maximum in the first 1.4 cm of the column, with a value of 4286 ___ 157 mg U/kg Fe (measured by ICP-MS). It has to be noted that the iron content of this sample was only 40% by weight instead of the initial 70%. Therefore, not only the maximum capacity of the iron to immobilise uranium has to be taken into account when designing a PRB but also the loss of reactive material with time. Samples with high uranium content were analysed with an Environmental Scanning Electron Microscope (ESEM) but it was not possible to identify any uranium mineral such as uraninite. It was presumed that either the uranium was amorphous or it was sorbed rather than precipitated.

C. Experiments with radiotracers A knowledge of the uranium behaviour in contact with elemental iron is needed if PRBs using iron are to be designed and their operation life estimated. Therefore, 237U was used as a radioindicator in column experiments to track the movement of uranium through the column without disturbing the system by taking samples or dismantling the apparatus. Soon after the detection of radioactivity, radioindicators (also called radiotracers) have been utilised for the investigation and analysis of processes and behaviour of material


87

components in various fields of application, e.g. medicine, chemistry, physics, material science, etc. (Gardner and Ely, 1967; Schulze et al., 1993; Gardner et al., 1997). Results from the experiments with such radiotracers, as described in the following sections, will be used to gain a better understanding of the uranium uptake capacity and thus of the longterm performance of PRBs using elemental iron or hydroxyapatite as reactive material.

1. Materials and methods 1.1. Experimental set-up Experiments with elemental iron Two test columns made of black polyethylene with a wall thickness of 12 mm, an inner diameter of 100 mm and a height of 540 mm were filled with sand/iron reactive material mixtures (mixing ratio 80:20 and 90/10, respectively). A uranium solution containing 237U without other minerals or salts was percolated from bottom to top through the column at different flow rates. The porosity of the iron/sand mixture was approximately 30%. The pH of the feed solution was neutral, but the pH of the column effluent was between 8 and 9 throughout the experiment (Reactions (4.3) and (4.4)). The experimental parameters of the two experiments performed are summarised in Table 4.3.

Experiments with hydroxyapatite A test column with the same dimensions as described above for the experiments with elemental iron was used for column experiments with hydroxyapatite. The column was filled with a 2O-cm layer of sand/HAP reactive material mixture (sand:HAP ratio 90:10), and with 15 cm sand layers above (outflow) and below (inflow) the mixture. The hydroxyapatite used in the experiment was a fine white powder with a grain size of 99% < 0.4 mm (see section B 1.2 Materials), the sand used was regular filter

Table 4.3. Experimental details of the column experiments with elemental iron using a 237U radiotracer. Experiment l Uranium concentration Flow rate Feed of uranium to the column Length of reactive zone in test column Mass of reactive material (sand and iron) Sand:iron mass ratio in reactive material

2.8 mg/l 0.276 l/h 0.773 mg/h 0.5 m 6.55 kg 80:20

Experiment 2 Uranium concentration Flow rate Feed of uranium to the column Length of reactive zone in test column Mass of reactive material (sand and iron) Sand:iron mass ratio in reactive material

3.11 mg/l 0.286 l/h 0.89 mg/h 0.15m 1.92 kg 90:10


88

sand (0.25-1.0 mm). A uranyl solution containing 3.1 mg/l of U (including the 237U) was percolated through the column at a flow rate of 0.286 l/h. The porosity of the reactive mixture was approximately 30% resulting in a flow velocity of 5.8 pore volumes per day. No attempt was made to exclude carbon dioxide from the experiment. The experimental parameters are listed in Table 4.4.

1.2. Preparation of activated uranium

237U can be produced by exposure of 238U to high energy (> 5 MeV) photon radiation from an electron accelerator 238U --1-hv

~

237U+ n

(4.12)

The half-life of 237U is 6.75 days. This isotope emits both ~/-rays and X-rays with the following photon emission energies: 59.5 keV (~/), 97 keV (X), 101 keV (X), 114 keV (X), 118 keV (X) and 208 keV (~/), making it easily detectable, e.g. using a scintillation counter. The isotope was produced by the photoneutron reaction on natural uranium in B AM' s linear electron accelerator. A sample of 1.1467 g U308 (0.9976 g of natural uranium contains approximately 0.9904 g of the isotope 238U) was irradiated with 30 MeV bremsstrahlung. The total exposure period was l OO min. Thereafter, the oxide was dissolved in concentrated nitric acid. After dilution with distilled water the concentration was 1.88 mg total uranium per ml. An aliquot of 150 ml was dissolved homogeneously in 1OO l of demineralised water resulting in a total uranium concentration for the batch experiment of 2.81 mg/l. The fraction of 238U converted to 237U according to Reaction (4.12) is not known. However, this ratio is not needed to perform quantitative calculations. The 237U signal is always proportional to the uranium concentration flowing through the column if radioactive decay is accounted for by a suitable correction.

1.3. Measurement and analytical procedure The solution containing the 237U tracer was percolated through the test columns containing the reactive material/sand mixtures. A ~/-detector was scanned vertically along the column recording the number of counts. The experimental set-up is displayed in Figure 4.9. A correction for radioactive decay of the nuclide was performed, but a correction for

Table 4.4. Experimental details of column experiments with hydroxyapatite (HAP) using a 237U radiotracer. Uranium concentration Flow rate Feed of uranium to the column Length of reactive zone in test column Mass of reactive material (sand and HAP) Sand:HAP mass ratio in reactive material

3.1 mg/l 0.286 l/h 0.89 mg/h 0.2 m 2.9 kg 90:10


89

Discharge

I Pulse counter I Test column filled with sand and iron

d collimator

Sodium iodide ~,-detector

Uranium solution reservoir

Figure 4.9. Schematicrepresentation of the experimental set-up.

absorption of photon radiation by the polyethylene pipe was unnecessary because it was found to be negligible. Samples of the column effluent were taken in 1-I batches at regular intervals (typically twice a day). The samples were collected in Marinelli beakers and measured using a high resolution ~/-spectrometer containing a large-volume germanium detector. Hereby, the complete spectra of all detectable radiocomponents were measured simultaneously (see Figure 4.10). Thus, radionuclides other than 237U, primarily fission products generated unavoidably during uranium photoactivation, could be traced. The photon emission spectrum of the test solution 3 days after uranium activation is shown in Figure 4. lO. It is obvious that most of the signals are due to photofission products whose activities (represented approximately by the height of the respective spectral lines) are low against that of 237U.The activities of all radioactive products detected 6 days after photoactivation are compiled in Table 4.5. Although the activities of the photofission 105

.... I .... ~'----V-,,,~ 237 u j~lllj,,~,,,,~3JJt I 132Te

l ....

i ....

1321

104 0

03

E 102 z 101 500

1000 Energy (keY)

Figure 4.10. Photonemission spectrum of the test solution.

1500

i

90


Table 4.5. Radionuclides produced through photoactivation of natural uranium. Nuclide

237U 93y 95Zr 97Zr 99M0

99mTc lO3Ru 127Sb 131Te 132Te

131I 132I 133I 140Ba 140La

141Ce

143Ce 153Sm

Half-life 6.75 d 10.1 h 64 d 16.8 h 66 h 66 h 39.3 d 3.85 d 30 h 76.3 h 8.02 d 76 h 21 h 12.8 d 40.3 h 32.5 d 33 h 46.75 h

Activity (Bq/l)

Count rate (s-l)

Partition of the total count rate (%)

125,914 128 661 336 3247 3113 999 81 1000 3161 2245 2313 831 2774 2488 1460 1497 1141

1320 0.9 1.2 1.0 1.4 5.4 4.0 0.5 9.9 12.3 5.1 9.3 1.1 2.2 1.8 7.3 1.8 2.0

95.16 0.06 0.09 0.07 0.10 0.39 0.29 0.04 0.71 0.89 0.37 0.67 0.08 0.16 0.13 0.53 0.13 0.14

products were low in comparison with 237U, they were monitored throughout the experiment as their migration velocities within the column would be different from 237U (due to their different chemistry). Hence, pulses counted by a sodium iodide (NaI) scanning probe can be attributed to 237Uwithout significant error. This is all the more true because the counting efficiency of the NaI detector favours the registration of low energies such as those emitted by 237U,but is inefficient in the energy region emitted by the fission products.

2. Experimental results 2.1. Elemental iron columns Immediately upon starting the experiment, low 237Uconcentrations were measured in the column effluent, but the effluent 237U concentration dropped below the detection limit within a few hours. It is assumed that the iron corrosion reaction has to get started before the uranyl can be reduced. The activity as a function of column length was first measured after 4 days. For Experiment 1, Figure 4.11 shows the activities measured at different time intervals in the column as a function of column height. The experiment was terminated after 860 h due to low counting rates after 5.3 half-lives of the uranium isotope (the uncorrected pulse count after 860 h was 5800 counts per minute, which is very low). The individual curves start with a positive slope, reach a maximum and finally decline towards low values. The shape of the curves is consistent with the post-test


91

250000 200000 /

g.9~_15oooo E

1: 96h

~

,

2:120

~~'N~\ \ \

a~ 100000 [

~

h

3: 168h 4:216 h 5:336 h 6:480 h 7:642 h

\ ~, \

13_

50000

0 j

,

0

2

,

4

i

,

6 Position (cm)

'

8

10

12

Figure 4.11. Activities measured in the iron column after different time intervals of the experiment (Experiment l). observations of Morrison et al. (2001) who found low concentrations at the inlet of a 1.2-m test column and the m a x i m u m concentrations at a distance of between 20 and 40 m m from the inlet. The recorded number of counts is proportional to the 237Uconcentration within the column. If the radioactive decay of the isotope is accounted for by a suitable correction, this figure is also proportional to the total uranium concentration because a definite 237U/238U ratio was initially set by the irradiation in the linear electron accelerator. If the feed concentration and the flow rate are stable over the duration of the experiment, a constant uranium flux is loaded onto the column. The areas under the curves displayed in Figure 4.11 representing the total uranium flux plotted against time should result in a straight line. Figure 4.12 clearly shows that this is the case.

2000000 1800000 1600000 = 1400000 o 1200000 1000000 E 800000 600000 400000 200000 0

J

E

O

O

ill 0

200

400

600

800

1000

Time (h) Figure 4.12. Integral number of counts as a function of time (Experiment 1). The product of flow rate, concentration and time yields the mass loaded onto the column.

92


The mass of accumulated uranium in the column at a given time can therefore be calculated as the product of flow rate, concentration, and time (see Table 4.3). To understand the long-term behaviour of PRBs, it is important to know the rate at which a contaminant front migrates through a laboratory test column, so that the contaminant breakthrough time for a PRB can be predicted. The radiotracer method enables the contamination front to be tracked as a function of time without interfering with the system. The most advanced position of the contaminant front in the column was arbitrarily defined as the highest position along the column where the recorded counts displayed in Figure 4.11 exceeded 20,000 per minute. The movement of the contaminant front through the column is plotted against time in Figure 4.13. A linear regression curve can be fitted to the data which suggests a linear passage of the contamination front through the column. These test results can therefore be used directly to predict barrier working-life, which simplifies the calculation of the amount of reactive material required for a particular situation. The migration velocity of the contamination front (slope in Figure 4.13) in the experiment was estimated to be ( 9 . 6 _ 0 . 4 ) x 10 -3 cm/h. From the experimental parameters shown in Table 4.3, a stoichiometric ratio between uranium and iron (see Reaction (4.1)) of 1:(1390 ___ 62) was calculated ( 0 . 7 7 6 m g - 3 . 2 6 x 10 -6 mol uranium reacted with 9 . 6 • 10 -3 cm equivalent to 0.25 g Fe = 4.5 X 10 -3 mol Fe). This suggests that 1390 moles of iron were necessary to remove 1 mole of uranium from the solution in the test columns operated with elemental iron reactive matrix. This result is coherent with the fact that both reactions and sorption can take place on the surface of the iron particles only. Therefore, most of the iron contained in a barrier does not take part in the uranium removal and a large amount of excess iron is needed. In the presence of excess elemental iron Reaction (4.1) is of pseudo-first-order kinetics (Cantrell et al., 1995; Gu et al., 1998) and can be described by Equation (4.13) d(U022+)/dt -

- k(U022+)

(4.13)

When the natural logarithm of the uranium concentrations measured at positions above the maxima in Figure 4.11 is plotted versus time a straight line is obtained. The slope of 9 8-

E 7O Lt - ' 6 -

5 E 4"E 3 O

t~

21

J

-

0"0

m

9 _I I

i

i

i

i

200

400

600

800

1000

Time (hours)

Figure 4.13. Movementof the contaminant front through the column as measured by the 237Uradiotracer (Experiment 1).


93

this straight line yields the rate constant of Reaction (4.1) In (UO 2+) -- - k At + const.

(4.14)

Based on this analysis the rate constant equals (1.1 + 0.09) x 10 -3 s -1. This value is in good agreement with the results of 6.9 x 10 -4 s-1 given by Cantrell et al. (1995) and 2-6.5 x 10 -3 s -1 given by Gu et al. (1998). However, the experimental results do not allow any conclusion with respect to the exact mechanism of removal, i.e. whether it is sorption, reduction plus precipitation or another mechanism. It can be shown that a steady-state uranium concentration per spatial element is reached rapidly. Figure 4.14 displays the number of counts recorded with the ~/-detector, which is proportional to the concentration of uranium measured at the column inlet as a function of time. After a certain period of time, a steady state is reached and uranium is not enriched further in the observed spatial element at the inlet and will move further through the column. This behaviour may provide some evidence against a mechanism of adsorption on iron oxyhydroxides, at least for short-term periods as in the present study ( < 40 days). The highest concentration of these corrosion products is expected at the column inlet where most of the oxygen consumption of the feed solution takes place. If sorption onto these substances would be responsible for uranium accumulation, uranium concentration should be greater at the inlet. However, the front of highest uranium concentration moves upwards through the column. A second experiment was conducted to verify these findings (Table 4.3, Fig. 4.15). The same experimental set-up as described for Experiment 1 was used, except for the iron:sand mass ratio that was 10:90. Figure 4.16 shows that the uranium fed-rate estimated from area under the curves in Figure 4.15 is constant, as would be expected from the constant flow-rate and influent concentration. The slope is in good agreement with the first experiment with similar flow conditions and uranium concentration. The initial velocity of the pollutant through the column is high because it passes first through a sand layer (Fig. 4.17). When the reactive layer is entered the velocity decreases to a constant small value. The velocity is 19.6 x 10 -3 cm/h, almost twice that in Experiment 1 where there was double the amount of iron.

106 105 E9

i - i - l l - - I

....

04

103 t'-

o=

0

102

101 100

i

0

200

400 600 Time (hours)

i

800

1000

Figure 4.14. Number of counts (i.e. uranium concentration) measured at column inlet (Experiment l).


94 250000

39

200000

1-

18h 24h 42h 66h - 90h - 162h - 235 h -331 h 9-429 h -

t-

~9

150000

~

,r v

if)

o

o

100000

t

50000

1-}

-5

0

17"-.

A_ r

0'

~.

v

..

5

I-

-"-

,~

,"

10

~

,m

,~

~

_

..

,',

m

15

I

20

25

H e i g h t (cm)

Figure 4.15. Uranium activities measured in the iron column of Experiment 2 after different time intervals. The stoichiometric factor was estimated to 1" 1200 which is in good agreement with Experiment 1. The maximum concentration of uranium was 3550 mg/kg, slightly higher than in E x p e r i m e n t 1. The pseudo-first-order rate constant was estimated to 1.39 x 10 -3 s -1, also slightly higher compared to Experiment 1.

Long-term performance of iron in column experiments Calculations for Experiment 1 yielded a uranium breakthrough time of 5200 h for the 50 cm column with an approximate 3000 mg/kg final uranium concentration on the iron particles. This result is consistent with the observation of Morrison et al. (2001) who found a maximum uranium concentration of 3400 mg per kg reactive material in a field column. 1400000 1200000 1000000 800000 600000 . . . . .

400000 200000 0

0

J

R2 = 0.9957

I

100

_ _ _

I

200 300 Time (hours)

Figure 4.16. Integral number of counts as a function of time.

400

500


95

16 1412-

~" 1oo ,..4 tO

:~ O

n

y = 0 . 0 1 9 6 x + 5.6681

8-

R2 = 0.9756

6---

1

4--

o -2 o

500

Time (hours) Figure 4.17. Movement of the contaminant front through the column of Experiment 2 as measured by 237U as a radioindicator (open symbols not used for calculation of the trendline).

In that work, chemical analysis was used to measure the maximum uranium concentration on the iron particles. In our investigation only the 237Usignal, which is proportional to the total uranium concentration, was used to calculate the final uranium concentration. A uranium mass fraction of 8% was found in iron nodules scavenging uranium from groundwater under natural conditions at the Koogarra uranium deposit in Australia (Sato et al., 1997). However, this accumulation has been formed during geological periods of time. Flow conditions chosen in this experiment represent an acceleration compared to natural conditions. The velocity in our column was around 5 pore volumes per day, i.e. 2.5 m/day, which is approximately five times the natural groundwater velocity. A PRB scaled-up from our experiments would persist for more than 3 years with the same pollutant concentration at natural velocities. The period of proper functioning could even be longer. Morrison et al. (2001) found that the removal efficiency increased again after a breakthrough in a column experiment after 3000 pore volumes when the flow velocity was decreased by a factor of 10. The removal capacity also depends on the grain size of iron used as reactive material. Mallants et al. (2001) found differences of a factor of 30 in the removal capacities between fine and coarse iron. A further limitation of the Fe ~ technology is the decrease of performance by corrosion and precipitation of minerals resulting in a loss of permeability. Groundwater chemistry is affected by dissolved oxygen and by the presence of anions such as sulphate and carbonate. These effects are discussed in detail elsewhere (Gu et al., 1999; Mackenzie et al., 1999) and have not been investigated in this study.

2.2. Hydroxyapatite columns The radiotracer method was also applied to investigate the behaviour of uranium in PRB systems with HAP as a reactive material. Although the uranium attenuation mechanism itself cannot be fully elucidated with the applied method, some useful data for the design of PRBs have been attained.

96

F.-G. Simon, V. Biermann 250000

1-

200000

'E" p__ 9 150000 ,,, ~= 100000

'

"-"

9

2-

l

- - -

4- 93h

+ _ -/7/7-~ / / |

Oh 13h

5-162h _ __ 6-233 h 7-285h 8-327h

///-~ \\ / L / / / ~6 " ~ \ \ //~" \ \\\ .

.

.

.

.

50000

0

0

5

10

15

20

Column height (cm)

Figure 4.18. Uranium isotope activities measured in the HAP column after different time intervals.

In the HAP column experiment, the activity of 237U was measured as a function of column length for the first time 13 h after experiment start. The experiment was finished after 451 h, i.e. a period of nearly 3 half-lives of the isotope. The uncorrected activity at that time had diminished to 14.5% of the initial value. Figure 4.18 shows the corrected pulse count along the column recorded after different time periods (9 of 17 curves recorded are displayed). A plot of the areas below the curves in Figure 4.18 against time results in a straight line (see Figure 4.19). The mass of uranium retained on the column can be calculated as the product of uranium concentration, time, and flow rate, i.e. 0.89 mg U per hour in the present experiment. 1000000 800000 600000 400000 200000 "0

100 '

200 ' 3 0 Time (hours)

46 0

500

Figure 4.19. Integral number of counts as a function of time. The product of flow rate, concentration and time yields the mass loaded onto the column.


97

16 .~ E

14 10

.,l:::

~)

8

E

6

"5 0

4

.i

..S

12

.T= o')

2

---i~

ll"

._IL

O

i

100

i

200

i

300

i

400

500

Time (hours)

Figure 4.20. Movement of the contaminant front through the HAP column as measured by 237U as a

radioindicator.

Figure 4.18 clearly shows that the uranium front moves forward through the column. To quantify the velocity of the movement of the contamination front, the position was arbitrarily set to that point where less than 5000 counts (which is approximately 2.5% of the maximum recorded value) have been recorded. Plotting this position against time of the different measurements should result in a straight line if the HAP is distributed homogeneously within the column. This is displayed in Figure 4.20. After a rapid increase in the first 45 h a linear behaviour can be observed (see first point in Figure 4.20 recorded after 13.4 h). The slope of the curve (without the first point) is the migration velocity of the contamination front and was estimated to be (21.2 +__0.8) • 10 -3 cm/h. The breakthrough of the contamination through the layer of reactive material in the column and the stoichiometric factor of the reaction of uranium with HAP can be calculated with these data. A breakthrough would have occurred after 944 h. By that time a total mass of (840 _ 32) mg uranium would have been loaded on the column resulting in a maximum concentration of (2916 _ 112) mg U/kg HAP. HAP is present in large excess so that the kinetics of uranium removal caused by HAP (Reactions (4.7)-(4.11)) is of pseudo-first-order which can be expressed by equation (4.13). When the natural logarithm of the uranium concentrations measured at positions above the maxima in Figure 4.18 is plotted versus time a straight line is obtained. The slope of this straight line gives the pseudo-first-order rate constant k. According to this analysis, the apparent rate constant equals 1.1 + 0.1 x 10 -3 S - 1 . From this value the half-life of uranium removal in the column can be calculated to 630 s. With a thickness of 20 cm for the layer containing the reactive material and a flow velocity of 3.4 x 10 -3 cngs the concentration of uranium after passing the reactive layer should be as low as after 9.4 halflives, i.e. ( 0. 5) 9.4 = 0.148%. A chemical analysis of the column leachate indicates that this may be the case as the measured concentrations are in the range of a few p.g/l. It has been shown that the rate of groundwater pollutant removal is a function of the surface area of the reactive material (Johnson et al., 1996). No attempt was made in the present work to quantify the influence of the surface, i.e. the grain size of the reactive material, on the reactivity.


98

It is possible to extrapolate the maximum uranium concentration on the HAP from the experimental data. Breakthrough of the pollutant through the column was calculated to occur after 944 h with approximately 2900 mg/kg of final uranium concentration on the HAP particles. This value results in a distribution coefficient of Kd = 2900/3.1 = 935.5 mg l/(kg mg). From the distribution coefficient Kd the retardation factor R can be calculated (Yong et al., 1992) as

R = 1 + (p/O)Kd

(4.16)

with p representing the bulk density and O the porosity, p/O can be assumed to be 6 kg/l for average conditions (Appelo and Postma, 1993). Thus, the retardation R of the 10%/ 90% HAP/sand mixture is R = 1 + ( 6 X 2 9 0 0 X 1 0 % / 3 . 1 ) = 562. The retardation calculated by dividing the velocity of the uranium solution by the velocity of the contamination front yields R = 3.4 x 10 -3 x 3600/21.2 X 10 .3 = 577 showing a good agreement. The stoichiometric factor of uranium removal as per Reactions (4.7)-(4.11) was estimated to 1:(487 _+ 19). The column experiments with elemental iron as reactive material resulted in a stoichiometric factor of 1:(1390 _+ 62). As both reductive precipitation and sorption are processes that occur on surfaces the difference in stoichiometric factors between HAP and Fe ~ by a factor of nearly three can be explained by the different grain sizes of both materials. In order to yield the same surface area a 2.5 times bigger volume of Fe ~ with a grain size of about 1 mm would be necessary in comparison to HAP particles of 0.4 mm diameter.

Long-term performance of HAP The results of the experiments showed that uranium can be removed from groundwater using hydroxyapatite as reactive material. This is in agreement with the promising results with HAP demonstrated at Fry Canyon in Utah (Naftz et al., 1999). However, it was suggested in a recent publication that PRBs using bone char as reactive material are prone to failure (Fuller et al., 2002b). In that work different apatite materials have been evaluated for their use in PRB systems for the remediation of uranium-contaminated groundwater. Breakthrough curves have been measured for several phosphate rock and bone meal samples. After the breakthrough, the columns were further operated with uranium-free artificial groundwater showing the reversibility of uranium (VI) sorption. Breakthrough for the material used at Fry Canyon occurred after approximately 600 ml groundwater/g bone char and 600 ppm of uranium sorbed. In our investigation no breakthrough occurred after 450 ml uranium solution/g HAP and 1400 ppm uranium sorbed. It is possible that the food-grade HAP used in the present investigation has a higher reactivity than the material used at Fry Canyon (Moore et al., 2001). In contrast to the reaction of lead with HAP where the sparingly soluble mineral hydroxypyromorphite Pblo(PO4)6(OH)2 is formed after the dissolution of HAP (Ma et al., 1993; Chen and Wright, 1997a), uranium removal might occur via sorption on HAP surfaces. In the non-mined Coles Hill uranium deposit phosphate mineral precipitation has been observed (Jerden and Sinha, 2003). Such a transformation was described by Sowder et al. (1996) for schoepite ((UO2)(OH)2). However, Jerden and Sinha (2003) stated that the formation of the autunite group minerals requires long time periods.


99

Locock and Burns (2003) reported the involvement of bacteria in the precipitation of autunite-group minerals. Geochemical simulation with the modelling software Phreeqc (Parkhurst and Appelo, 1999) indicates that uranium (VI) may be precipitated by HAP. Saturation indices >O were calculated for schoepite and the sparingly soluble uranium-phosphorus compounds autunite (see Reaction (4.7)), chernikovite (see Reaction (4.8)) and (UO2)3(PO4)2. The Minteq database from the computer code MinteqA2 was used for the calculations (Allison et al., 1991). The species UO2(HPO4) 2- of the speciation model proposed in an early work of Langmuir (1978) was removed from the database, as agreement now exists that this species should be disregarded in geochemical modelling of uranium (Sandino and Bruno, 1992; Langmuir, 1997). Which mineral's formation is predicted by a geochemical model strongly depends on the minerals' solubility products (Ksp) in the data base. However, the values differ within a broad range. For chernikovite, values from - 45.46 (Van Haverbeke et al., 1996) to - 5 0 . 0 3 (Johnson, 2000) have been obtained for log Ksp. In the review of Grenthe et al. (1992), log K s p - - 4 8 . 4 is recommended. The formation of schoepite is strongly dependent on the pH (Sandino and Bruno, 1992; Giammar, 2001). Although the results from geochemical modelling indicate that uranium may be removed by precipitation after contact with HAP, ion exchange or sorption processes cannot be ruled out because no attempt was undertaken to identify solid phases on the reactive material from the test column. In the batch experiments performed by Fuller et al. (2OO2a), chernikovite could not be observed until more than 7000 ppm of uranium were sorbed. Autunite was observed at even higher uranium-phosphate ratios. At lower uranium concentrations, the formation of an inner-sphere complex of U(VI) to HAP surfaces was suggested. The mechanism of heavy metal removal by HAP is different for lead, cadmium and zinc. Whereas lead removal is almost independent from pH indicating precipitation, the solubility of cadmium and zinc increases at pH values at and below 6 which may be explained by ion exchange or adsorption (Chen and Wright, 1997a,b). It is conceivable that uranium behaves like zinc and cadmium rather than lead. For antimony an ion exchange mechanism was proposed (Leyva et al., 2001).

D. Conclusions

Using the radiotracer method the concentration of uranium retained in the columns can easily be measured without interfering with the flow regime or dismantling the apparatus. Investigation of the contaminant's spatial distribution at any given time in the reactive media would be difficult without a radioactive tracer. This method provides a powerful tool to elucidate uranium retardation within the reactive media while maintaining undisturbed flow conditions. The results obtained using this method will help provide further information about processes within barriers and the quantity of reactive material needed to treat a given amount of contaminated water with known chemical properties. Conclusions regarding the respective removal mechanism cannot be easily derived from the results. For the reaction of uranium with iron some evidence was found against a mechanism of adsorption on iron oxyhydroxides, at least for short-term periods. Whether sorption of uranium by HAP occurs via a sequence as displayed in Reactions

1OO


(4.10) or (4.11) strongly depends on the speciation of uranium as a function of pH. The uranyl ion UO 2+ is stable at low pH values only. When phosphate and carbonate are present the speciation is dominated by carbonato complexes (UO2CO3(aq),U02(C03) 2and U02(C03) 4-) at a pH--> 8, while soluble phosphato complexes (UO2H2PO4+, UO2HPO4(aq), UO2PO4, etc.) are formed between pH 4 and 7 (Giammar, 2001). Therefore, ternary complexes including additional anions like carbonate or phosphate are also conceivable. Structures have been suggested for such ternary complexes for the sorption of uranium on hematite surfaces (Bargar et al., 2000) and on HAP (Bargar et al., 2002). The formation of stable sorption complexes or mineral phases is essential to obtain the desired behaviour from a PRB system. Evidence exists that under normal conditions uranium remains immobile in HAP barriers (Giammar, 2001). The data presented in this study facilitate safe PRB design by making it possible to calculate the amount of reactive material necessary for the removal of the contaminants. However, lessons learnt from the demonstration project at Fry Canyon show that characterisation of both the site and the groundwater flow regime is very important. If any of the groundwater by-passes the reactive medium, the removal efficiency is low. The present production cost of HAP (which might be reduced by new production methods) is 500 US$ per ton for Apatite II (Conca and Wright, 2003). This is in the same order of magnitude as for granular iron (350 US$ according to EnviroMetal Technologies Inc.). Lower prices for HAP are feasible if HAP containing waste materials are used (e.g. fishbone or phosphatic clay; Singh et al., 2001). One advantage of apatite compared to elemental iron is its extremely high uranium retaining capacity of up to 30% whereas the maximum observed load in elemental iron was around 1000 mg/kg after 14 months of in situ treatment (Bostick et al., 2000).

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Simon, F.G., Meggyes, T., McDonald, C., 2002. Advanced Groundwater Remediation - Active and Passive Technologies. Thomas Telford, London. Singer, E., Matucha, M., 1962. Erfahrungen mit der Bestimmung von Uran in Erzen und Gesteinen mit Arsenazo III. Zur Analytischen Chemie 191,248-253. Singh, S.P., Ma, Q.Y., Harris, W.G., 2001. Heavy metal interactions with phosphatic clay: sorption and desorption behaviour. J. Environ. Qual. 30 (6), 1961-1968. Sivavec, T.M., Homey, D.P., 1995. Reductive Dechlorination of Chlorinated Ethenes by Iron Metal, 209th American Chemical Society National Meeting, Anaheim, CA, Division of Environmental Chemistry, Conference Proceedings, Vol. 35, pp. 695-698. Sowder, A.G., Clark, S.B., Fjeld, R.A., 1996. The effect of silica and phosphate on the transformation of schoepite to becquerelite and other uranyl phases. Radiochim. Acta 74, 45-49. Van Haverbeke, L., Vochten, R., Van Springel, K., 1996. Solubility and spectrochemical characteristics of synthetic chernikovite and meta-ankoleite. Mineral. Mag. 60, 759-766. Wu, L., Forsling, W., Schindler, P.W., 1991. Surface complexation of calcium minerals in aqueous solution, 1. Surface protonation at fluorapatite-water interfaces. J. Colloid Interface Sci. 147 (1), 178-185. Yong, R.N., Mohamed, A.M.O., Warkentin, B.P., 1992. Principles of Contaminant Transport in Soils. Development in Geotechnical Engineering, Vol. 73. Elsevier, Amsterdam, 327 p.


104

Appendix 4A. Results of geochemical simulation of the column experiments feed solution with PHREEQC (based on the data given in Table 4.1) Description

of s o l u t i o n .

pH pe A c t i v i t y of w a t e r Ionic strength M a s s of w a t e r (kg) T o t a l a l k a l i n i t y (eq/kg) T o t a l C02 (mol/kg) T e m p e r a t u r e (~ E l e c t r i c a l b a l a n c e (eq) P e r c e n t error, i00 X (Cat - IAnl)/(Cat + IAnl) Iterations Total H Total 0

7.000 4.000 0.999 3. 945 X 10 -2 1.000 i. 036 X 10 -2 1 . 2 1 7 X 10 -2

-

-

25.000 1 . 8 4 2 X 10 -3

--3.52 ii 1.110264X102

5. 5 5 7 9 0 5 X i0

,0..

D i s t r i b u t i o n o f species.

Species ~.

Molal ity

Activity

1.214 x 1.158 X l o - ’ 5.551 X 10

1 . 0 0 4 X lO-~’ 1.000 x lo-’ 9.993 x 1 0 - I

Logrnolality

Logactivity

L o g gamma

.

OH-

H+ H2 0

C ( 4 ) 1.217 X HCO; HzC03 MgHCOf CaHCOl NaHC03 CaCO,

co4 -

MgC03

uoz(c03) $ UOz(CO3 1;NaCO;

u02c03 Ca 4.501 X Ca”+ CaS04 CaHCO: CaC03 CaOH+

-

6.916 6.936 1.744

-

6.998 7.000 0.000

0.082

~

- 0.064 0.000

3

s 9.871 1.849 2.027 1.706 4.559 9.179 7.747 6.345 1.953 1.595 8.425 3.028

X X

lo-’

lou3

x 10-~ X

loA4

x X

x

lo+

X

lo-&

x x X X

lo-’ lo-’

8.309XlO 1.866 X lop3 1.685 X 1.443 X l o p 4 4.600 X lo-‘ 9.262 X l O - “ 3.889 X 6.402 X 9 . 4 9 1 X lo-’ 8.892 X l o p 8 7.092 x 1 0 - ~ 3 . 0 5 6 X LO-’

2.006 2.733 - 3.693 - 3.768 - 4.341 - 5.037 - 5.111 - 5.198 - 5.709 - 5.797 - 6.074 - 7.519 ~

2.080 - 2.729 - 3.773 - 3.841 - 4.337 - 5.033 - 5.410 - 5.194 - 6.023 - 7.051 - 6.149 - 7.515

- 0.075

- 2.776

-

-

0.004 - 0.080 - 0.073 0.004 0.004 - 0.299 0.004 - 0.313 - 1.254 - 0.075 0.004

lo-’ 3.226 1.096 1.706 9.179 4.997

x 10-~ x 10-~ X X

x

loA4

lo-’

1.676 X 1.106 X lo-’ 1.443XlO 9.262 X l o p 6 4.227 X 10

-’

-

2.491

- 2.960 - 3.768 - 5.037 -

8.301

-

2.473

2.956 - 3.841 - 5.033 - 8.374

-

0.284 0.004 - 0.073 0.004 - 0.073

ci 3 . 3 6 4 x

c1uo2c1+ UCP+

a-. z

3.364 x 1 0 - ~ 3.440 x 3 . 1 4 9 X lo-*’

2.773X1V3 2.872 X 0.000

- 14.463 - 39.502

2.557 14.542 - 40.207 -

-

- 0.084 - 0.078 -

0.705

P

%’

2P

s (b

3

a

%. 9

(continued) Species

Mo 1a1i t y

Logmolality

Activity

Log activity

Log gamma

~~

H(0) 1.403X10-25 H2 K1.282X10 K+ KSO,

Mg4.452XlOP3 Mg2+ MEIS04 MgHCO?;

MgC03 MgOH+ N a 1 . 1 9 9 X lop2 Na+ NaS0; NaHC03 NaC0;

7 . 0 1 5 X lo-''

7.079 X

1.254 X l o r 2.793XlO

1.033 X 2.351X10-6

-

1 . 7 2 5 X lo-' 9.933 x 1.685 X 6.402XlO 2.831XlO

- 2.487 - 3.007

9.838 X 1.597 X 4.600 x 1 0 P 7.092 X l o - '

- 1.930 - 3.722

4

3.259 x l o r 3 9.844 X 2.027XlO 6 . 3 4 5 X lo-' 3 . 3 3 4 X lo-'

'

1.175 1.897 4.559 8.425

X lo-' X lo-' X

lop5

X

- 25.154

*

-

-

-

3.902 5.554

3.693 5.198 7.477

4.341

- 6.074

-

0.004

25.150

- 3.986 -

5.629

-

0.084

- 0.075

2.763 3.003 - 3.773 5.194 - 7.548

-

0.276 0.004

-

0.080

-

0.004 0.071

-

~

2.007

-

- 3.797 - 4.337

-

6.149

-

-

-

0.077 0.075 0.004 0.075

O ( 0 ) 0.000

0.000

0 2

0.000

-

3.238 X 1.106 X l o p 3 9.933 x l o p 4 1.597 X 2.351 X 3.142XlO 1.034 X

2.167 2.960 - 3.007 - 3.722 - 5.554 - 7.425 - 11.989

42.085

-

0.004

42.081

S(6) 9.077X10

so;

~

CaS04 MgS04 NaS0;

KS04 HSO;

uozso4

6.805 1.096 9.844 1.897 2.793 3.762 1.025

X 1 0 -3 X X

x 10-~ X X X

lops

*

-

- 2.490

-

2.956 - 3.003 - 3.797 - 5.629 - 7.503 - 11.985 ~

0.323 0.004 0.004 - 0.075 - 0.075 0.078 0.004 -

~

(continued) ~~~

Species

Molal i t y

~

Activity

Logactivity

Logmolality

Log gamma

~

2 . 0 5 3 X 10-I’ 5.992 X 1 0 -35 1.980X10 ”

9.975 X 6.047 X 9.622 x

0.000

0.000

9.305 X 3.226 X 1.427 x 1.134 X 1.133 x 5 . 9 9 2 X l O 35 1.980 x 1 . 8 4 4 X 103.149 X 1 0 43 0.000

7.769XlO-l8 3.256 X lopzo I.1 9 1 x 5.512 X 2.233 x 6.047 X 9.622 x 1.028 x 0.000

12.688 - 34.222 35.703 -

~

- 13.001 - 34.218 - 36.017

-

0.313 0.004 0.313

-

0.705

-

U ( 3 ) 0.000 u3

+

’’

4.557

x 1 0 -I4

1.953X10 1.595 X 3.028XlO * 4 . 7 1 4 X lo-’’ 1.285 X 1 . 0 2 5 X 1 0 l2

-

0.000 3.805 x 10-l~

’

9.491 X 1 0 8.892XlIY8 3.056 X lo-* 3.935 X 10-l’ 6.242 X 1.034X10-12

51.079

-

51.784

17.031 19.491 - 22.846 - 26.945 - 31.946 - 34.222 - 35.703 - 37.734 39.502 - 139.814

17.110 19.487 - 22.924 - 27.259 - 32.651 - 34.218 - 36.017 - 38.988 40.207 - 146.162

- 13.341

- 13.420

5.709 5.797 - 7.519 - 10.327 - 11.891 - 11.989

6.023 7.051 - 7.515 - 10.405 - 12.205 - 11.985

-

~

-

-

~

0.078 0.004 - 0.078 - 0.313 - 0.705 0.004 - 0.313 - 1.254 - 0.705 - 6.348 -

- 0.078

-

-

-

-

-

0.313 1.254 0.004 0.078 0.313 0.004

(continued)

c.

M o l a 1i t y

Activity

uo2 (SO4 12uo2c1+

2.053 x 3 . 4 4 0 X 10-l 5

(UO,)Z(OH)~+ ( UOz ) 3 (OH)

1.921 X 10 -lE 8.179 X

9.975 x 1 0 - l ~ 2.872 X 10 9.332 X 6 . 8 2 8 X lo-'*

Species

Logmolality -

12.688 14.463 15.717 17.087

Logactivity - 13.001

14.542 - 16,030 - 17.166 -

L o g gamma

- 0.313 - 0.078 -

0.313

- 0.078

Behaviour of uranium in elemental iron and hydroxyapatite reactive barriers Saturation

indices.

Phase

SI

Anhydrite

-

log

IAP

log

KT

0.63

-- 5 . 2 7

-

0.15

-- 8 . 1 9

-- 8 . 3 4

CaC03

Artinite

-- 6 . 5 4

3.06

9.60

MgC03

B_U02

-- 3 . 7 5

1.79

5.54

UO2(OH)2

Aragonite (OH) 2

4.64

Brucite

-- 5 . 5 6

11.24

16.79

Calcite

0.29

-- 8 . 1 9

-- 8 . 4 7

CH4(g)

CaS04

CaC03

-- 1 0 7 . 4 1

-- 4 0 . 1 0

CH4 C02

-- 1 . 2 5

-- 1 9 . 4 1

-- 1 8 . 1 6

0.64

-- 1 6 . 3 6

-- 1 7 . 0 0

CaMg(C03)2

Epsomite

-- 3 . 1 2

-- 5 . 2 6

-- 2 . 1 4

MgSO4:7H20

Gummite

-- 8 . 6 1

1.79

Gypsum

-- 0 . 4 2

-- 5 . 2 7

-- 2 2 . 0 5

-- 2 2 . 0 0

6.15

-

10.40 -

U03

4.85

CaS04

0.04

H2 NaCl

:2H20

Halite

-

4.56

1.58

Huntite

-- 2 . 7 4

-- 3 2 . 7 1

-- 2 9 . 9 7

Hydromagnesite

-- 1 2 . 6 9

-- 2 1 . 4 6

-- 8 . 7 7

Lime

-- 2 1 . 5 7

11.22

32.80

Magnesite

-- 0 . 1 4

-- 8 . 1 7

-- 8 . 0 3

MgC03

Mirabilite

-- 5 . 3 9

-

-

1.11

Na2SO4-lOH20

Natron

-- 8 . 1 2

--9.43

-- 1 . 3 1

Na2CO3:lOH20

Nesquehonite

-

-

-

MgC03

2.55

6.51 8.17

5.62

CaMg3(C03)4 Mgs(CO3)4(OH)2 CaO

02(g)

-- 3 9 . 1 2

44.00

83.12

02

Periclase

-- 1 0 . 2 7

11.24

21.51

MgO

Portlandite

-

11.22

22.68

Ca(OH)2

14.46

U02C03

11.45

Rutherfordine

-

3.15

Schoepite

-

3.61

Thenardite

-

6.32

Thermonatrite

-

17.61

-

1.79 -

6.50

-

:3H20

Mg(OH)2

-- 6 7 . 3 1

C 0 2 (g)

"Mg(OH)2

Dolomite

H2(g)

109

5.40

UO2(OH)2

0.18

Na2S04 Na2C03

-- 9 . 5 5

-- 9 . 4 2

0.13

U 3 0 8 (C)

-- 1 0 . 0 7

-- 1 6 . 6 1

-- 6 . 5 4

U308

U 4 0 9 (C)

-- 1 8 . 5 7

-- 5 8 . 8 2

-- 4 0 . 2 5

U409

U02 (am)

-- 1 1 . 9 2

-- 2 0 . 2 0

-- 8 . 2 8

U02

U03(C)

-- 5 . 9 2

1.79

7.72

UO3

Uraninite

-- 6 . 2 9

-- 2 0 . 2 0

-- 1 3 . 9 2

U02

93 H 2 0

:H20 :H20

:4H20



111

Chapter 5 Laboratory tests using natural groundwater Jdzsef B6hm, ,/~kos Debreczeni, Imre Gombk6t6, Franz-Georg Simon and Mihfily Cs6vfiri A. Introduction To estimate the operational life-time of a permeable reactive barrier (PRB) system, it is necessary first to estimate (i) how long the reactive material will retain its reactivity towards the target contaminant(s), (ii) the rate at which effects arising from major groundwater constituents will influence the performance of the system (especially its hydraulic permeability), and, if the barrier design allows replacement of the active material, (iii) how often that material is replaced. To verify such quantitative assumptions, laboratory column experiments are usually conducted on any reactive media intended for used in a PRB. These column tests usually form part of the feasibility studies for the PRB, and should be run under conditions that reflect the actual conditions at the remediation site. These tests differ from the simple experimental systems often used to investigate the fundamental processes by which a reactive material operates (i.e. the type of tests described in Chapters 3 and 4 for the retention of uranium by elemental iron and hydroxyapatite), as the aim is to replicate as accurately as possible the actual behaviour of a given system (e.g. elemental iron in the groundwater of a specific site). In this chapter, laboratory experiments using uranium-contaminated natural groundwater originating from the former uranium mining and processing site near Prcs, Hungary, are presented. Experiments in laboratory columns and floor-scale cells are described with special attention to the long-term behaviour of the reactive materials studied (elemental iron and hydroxyapatite).

B. Column experiments

l. Experimental set-up The experiments were carried out in 2.2 cm diameter glass columns (Fig. 5.1) filled to a height of 22 cm with the reactive material being tested (to give an overall volume of reactive material of 84 cm3). The contaminated groundwater was percolated through the columns using peristaltic pumps at a rate that gave a residence time of approximately 4 h. The composition of the effluent was analysed for uranium and some of the major

112

J. Brhm etal.

Figure 5.1. Glass column experiments using local groundwater.

constituents such as pH, specific conductivity, total dissolved solids (TDS), Ca, Mg, CI, HCO3, $04 and redox potential. Commercially available elemental iron (shredded grey cast iron supplied by Gotthard Mayer, Rheinfelden, Germany, with a typical particle size of 0.3-1.3 mm) and hydroxyapatite (HAP, supplied by Chemische Fabrik Budenheim CFB, Germany), as described in Chapter 3, were used as reactive materials in the experiments. Additionally, experiments were conducted on waste steel fibres (Fig. 5.2) that are a by-product of tyre recycling at a facility in Prcs (i.e. a low cost material). Uranium-contaminated natural groundwater from the former uranium mining and processing site near Prcs, Hungary, was used for the experiments (see Chapter 9 for details of the Prcs site). This was taken from three different locations at that site: from monitoring well Hb-01/1 downstream of waste rock pile III, from monitoring well TV-5 at the former uranium ore milling site and seepage water from the tailings pond area. The average compositions of these groundwater samples are shown in Table 5.1. The column tests conducted on groundwater pumped from a monitoring well (Hb-Ol/l) are of special interest because this well is close to the site of the pilot scale PRB (Chapter 10). The groundwater at monitoring well Hb-01/l was highly contaminated with uranium with concentrations of up to 1000 lxg/l. The concentration of TDS was also high, with calcium,

113

Laboratory tests using natural groundwater

Figure 5.2. Waste steel fibres from tyre recycling.

magnesium, sodium, bicarbonate and sulphate being the major constituents. The groundwater at monitoring well TV-5 was characterised by a high pH and high bicarbonate content and uranium concentrations of around 1500 txg/l. The seepage water from the tailings ponds showed the highest mineralisation accompanied with a high uranium concentration of > 2000 ~g/l (Table 5.1). The experiments were conducted in a laboratory on the P6cs site, close to the sites from where the water has been pumped. While most experimental parameters were close to the in situ conditions, some parameters such as dissolved oxygen content and temperature changed significantly during transfer of the pumped groundwater to the laboratory. The initial oxygen content of the local groundwater was between 0.2 and 0.5 mg/l of 02, but increased in the laboratory to approximately 5 mg/l of 02. The groundwater temperature in situ was approximately 11~ while it was around 20-24~ in the laboratory.

Table 5.1. Results of column experiments using elemental iron (shredded cast iron) percolated with uranium-contaminated groundwater: selected groundwater and column effluent constituents (average values, duration of experiments 60 days).

Parameter pH

Electrical U TDS conductivity (Ixg/l) (mg/l) (IxS/cm)

Ca Mg (mg/l) (mg/l)

H C O 3 804 CI Eh (mg/l) (mg/l) (mg/l) (mV)

Waste rock pile (groundwater from monitoring well Hb-01/1) Influent 7 . 1 1 1496 718 1030 168 53 474 Effluent 8.43 915 11.4 740 19 51 178 Former heap leaching pile (groundwater from monitoring well TV-5) Influent 8 . 6 4 1192 1500 888 17.6 6.2 585 Effluent 9 . 3 1 1113 7.7 771 6.4 4.6 305 Tailings ponds seepage water Influent 7.80 12,360 2320 17,080 400 1944 518 Effluent 7 . 2 3 11,253 18.3 12,604 380 1749 254

361

43 43

250 128

85 85

170 115

1455 1313

167 148

114

J. Brhm et al.

2. Experiments on shredded cast iron

The results of the column experiments with shredded cast iron are summarised in Table 5.1. The test duration was 2 months, and the initial hydraulic flow velocity was approximately 7 x 10 -6 rrds. The residence time of the groundwater in the columns was 4 - 5 h (the average value from the experiments was: 4.17 h), at a column porosity of about 40%. Uranium was removed very efficiently from the polluted water of all the three sites indicating that the elemental iron is indeed suitable for treating uranium-contaminated water from these sites (waste rock pile, former heap leaching site and tailings ponds). After passing through the columns, the general composition of the water changed significantly (Table 5.1). TDS of the water decreased by 15-25%. The hydraulic conductivity decreased continuously, most likely due to the accumulation of precipitates and the cementation process of fine iron particles.

3. Experiments on waste steel fibres The application of a waste material such as steel fibres from tyre recycling in groundwater remediation is economically attractive because the material is available at a low price and can to some extent exhibit similar properties as shredded cast iron. The bulk density of the material is low (--~ 1 kg/l), which may cause problems for some applications. After filling the glass column with the waste steel fibres the voids were filled with sand (0.63-1 mm). The duration of the column experiment was 12 months, and during this period a total of approximately 400 bed volumes of water (from well Hb-01/l) were passed through the column. The experimental results are summarised in Table 5.2. Like the experiments on shredded cast iron, uranium was removed very efficiently with the uranium concentration decreasing from 800 Ixg/l in the influent to 10-20 ~g/l in the effluent, in spite of the low mass of the iron in the column. The TDS of the groundwater decreased by only 16%, and the pH increase was 1.1 units, indicating less overall change in the groundwater than in the column experiments with shredded cast iron and the same influent water. Iron was present in the effluent at concentrations around 2 - 3 mg/l. The hydraulic conductivity has decreased only slightly, with a more rapid decrease in the final phase of the experiment. Generally, it can be concluded that waste steel fibres can be used for efficient uranium removal from groundwater.

Table 5.2. Results of column experiments on elemental iron (waste steel fibres) percolated with uranium-contaminated groundwater: selected groundwater and column effluent constituents (average values, duration of experiments 12 months). Parameter pH

Influenta Effluent a

Electrical U TDS Ca Mg HCO3 5 0 4 CI Fe Eh conductivity (txg/l)(mg/l)(mg/1) (mg/l) (mg/1) (mg/l) (mg/l) (mg/l) (mV) (txS/cm)

7.11 1496 8.20 1272

718 17

1030 168 867 82

53 60

474 367

361 421

Groundwater from monitoring well Hb-01/l downstream of waste rock pile 111.

43 50

250 2.6

250 234


115

4. Experiments on hydroxyapatite The hydroxyapatite used for the experiment was a fine powder. To obtain a good hydraulic conductivity, a mixture of HAP and sand (20%/80% by volume) was used in a column which was slightly larger than for the iron experiments (diameter 3.2 cm, volume 281 cm3). After approximately 150 bed volumes of groundwater (originating from monitoring well Hb-O 1/1) passed through the column, uranium breakthrough was detected (Fig. 5.3). The experiment was continued and after passing another approximately 250 bed volumes of water, a second HAP column was attached to the outlet of the first column and the experiment continued. In total, the experiment lasted more than 8 months. The coupling of the two columns resulted in an improvement of the uranium removal in the effluent of the second column, while uranium content in the effluent from the first column measured at a separate sampling point remained high. The experimental data obtained during the experiment are presented in Table 5.3. With the bulk densities of sand and HAP (1.5 and 0.5 kg/l, respectively) it can be calculated that the mass ratio of the HAP/sand mixture was 8%/92%. The removal capacity can therefore be estimated to be around 2000 mg U/kg HAP. The overall groundwater composition changed only slightly. A small decrease in the calcium concentration is likely due to an increase in pH. Only a minor amount of precipitation was observed, and therefore, clogging of the pores of the material mixture appears to be unlikely. In summary, HAP is effective at removing uranium from contaminated water. In conclusion of the laboratory column experiments, it can be stated that all the three reactive materials investigated are suitable for the removal of uranium from groundwater. When iron is used as the reactive material, precipitation of secondary minerals will

10000

1000

- rn

rnt~

03 "-1 v

d

c-

O L)

E

100

A.= V v

A v

t" t_

10

0

.= v

100

2+0

3+0

460

Bed volumes

5+0

6+0

700

Figure 5.3. Uranium concentration in the effluent of column experiments on HAP. After 400 bed volumes a second HAP column was coupled to the outlet of the first column. Concentration measured in the outlet of the first column after coupling of the two columns displayed with open squares (see text).

l 16

J. Bb'hm et al.

Table 5.3. Results of column experiments with hydroxyapatite (HAP) percolated with uraniumcontaminated groundwater: selected groundwater and column effluent constituents (average values, duration of experiments 6 months).

Parameter

pH

Electrical conductivity (IxS/cm)

U (Ixg/l)

Ca (mg/l)

Mg (mg/l)

HCO3

Influent a Effluent column 1 Effluent column 2

7.46 7.99 8.06

1726 1539 1463

913 < 10-900 b < 10-160 c

190 142 108

64 62 63

630 522 414

Groundwater from monitoring well Hb-01/1 downstream of waste rock pile III. b Uranium breakthrough (Fig. 5.3). c At the end of the experiment (Fig. 5.3). a

certainly take place, and care must be taken to avoid rapid loss of hydraulic conductivity in the reactive media.

C. Floor-scale tests

1. Design and operation Floor-scale tests have been performed in i m 3 test boxes. The test boxes were installed and operated in the premises of Mecsek Ore Environment Co. located on the former uranium ore milling site near Prcs, Hungary (see Chapter 9). Effluent from a mine water treatment plant located on that site was used in the experiments, since it was not possible to collect enough water from the monitoring wells located on the test site downstream of waste rock pile III (see Chapter 9) to continuously feed all three test boxes throughout the experiments. The effluent still has an elevated uranium concentration. The chemical analysis of this water in comparison to the water at the test site is displayed in Table 5.4. The composition of the treatment plant effluent is in general similar to that of the groundwater at the test site. The floor-scale tests were conducted in three identical boxes of l x l x l m 3 in size, which were purpose-built from stainless steel. Each box has three important features (Fig. 5.4). First, there is an overflow system at each end of the box that ensures the flow through the reactive media is driven gravitationally by a constant hydraulic gradient. This system created a flow regime that can be easily modelled, and where any changes can be observed and inspection is easy. Second, there are open water tanks before and after the reactive zone to facilitate a homogeneous flow through the treatment zone. The third feature is that the treatment zone consists of three layers: the reactive material and an up-gradient and down-gradient sand layer. This treatment zone is in the middle of the box separated from the water tanks by perforated stainless steel plates. The dimensions of the water tanks were 1 x 0.1 x l m and the space provided for the material filling was 1 x 0.8 x 1 m. Attempts were made to prevent preferential flow along the sidewalls of the box by constructing a labyrinth system along the box walls perpendicular to the water flow. The labyrinth consisted of corrugations that are 4 cm high with a distance of 5 cm between each of them. Bentonite was placed into the space between the corrugations. In contact with water


l 17

Table 5.4. Composition of water used in the floor-scale experiments and a typical groundwater sample taken from the test site downstream of waste rock pile III.

Na K Ca Mg CI $04 C03 HCO3 Carbonate hardness Total hardness pH U Electrical conductivity a

Experimental

Test site

Unit

275 5 180 108 119 870 10 741 34 50.4 7.2 850 2433

44 14 140 40 32 253 10 412 18.9 28.8 6.8 385 1061

mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l -~ a -~ mg/l IxS/cm

edH = Deutsche HS~egrade (German hardness, equivalent to 10 mg/l CaO).

this bentonite will have swollen, creating a seal at the sides of the box. Figure 5.5 shows photographs of the test boxes prior to filling them with reactive material. Floor-scale tests were conducted using elemental iron, steel fibres and hydroxyapatite as reactive media. During previous column tests using 100% elemental iron, corrosion products had aggregated within the pore space, reducing hydraulic permeability considerably. Therefore, the elemental iron was mixed with an equal volume of quartz sand in an attempt to achieve a more stable long-term hydraulic conductivity. Therefore, sand was used for the filtration layers and for mixing with the reactive material to achieve

/

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,o OoOo~

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9o o~'%, ,0- , , 0o2 oHiO...~ ~o_,:, o' 4 mm 2 - 4 mm 1 - 2 mm 0 . 5 - 1 mm 0.2-0.5 mm 0 . 1 - 0 . 2 mm 0.063-0.1 mm 0.045-0.063 mm 0.032-0.045 mm < 0.032 mm U (mg/kg) pH Electric conductivity (l~S/cm)

0.0-0.4

0.4-0.8

1.2-1.4

1.4-1.8

3.84 2.86 5.22 10.90 18.28 8.56 4.16 0.56 2.10 43.52

5.06 4.90 7.26 16.02 20.40 8.76 1.82 3.92 1.68 30.18

7.30 3.92 6.18 15.50 22.98 8.96 4.66 0.92 2.20 27.38

6.38 3.40 5.58 14.92 21.12 8.76 4.14 1.08 2.88 31.74

7 8.00 1020

8 7.87 1210

7 7.89 870

< 7 7.89 870

2.5-2.8

2.8-3.0

3.1-3.5

3.8-4.2

36.20 7.20 7.74 14.46 14.84 5.42 2.14 0.64 0.62 10.74

14.34 5.32 7.14 15.94 21.92 8.92 3.98 0.78 1.14 20.52

9.44 1.46 3.54 19.08 27.22 9.74 4.10 0.68 0.72 24.02

15.42 6.36 9.18 17.74 22.34 8.48 3.46 0.96 0.94 15.12

< 7 7.84 1240

< 7 8.12 560

< 7 8.12 530

< 7 8.12 740

Table 9.2. Results of the hydraulic tests (transmissivity T and permeability kf). T (me/s)

kf (m/s)

Hb-01/l Theis steptest a Cooper-Jacob steptest b Theis recovery test c

1.12 • 10 -4 8.34 x 10 -5 7.36 x 10- 5

2.81 X 10 -5 2.08 x 10 -5 1.22 X 10- 5

Pe-03 Theis recovery test c

9.17 x 10 -6

9.17 • 10 -6

Pe-Ol Theis recovery test c

1.00 x 10 -7

2.50 X 10 -7

a Cooper-Jacob steptest (variable discharge rate). The software code AquiferTest provides the ability to use water level vs. time data which were recorded during a variable rate or intermittent pumping test to determine the transmissivity and storativity. A time transformation is used to provide a congruent data set. b Theis steptest (confined). Theis solved the unsteady-state groundwater flow equation. For the variable rate pumping case, you can use water level vs. time data which were recorded during a variable rate or intermittent pumping test to determine the transmissivity and storativity. A time transformation is used to provide a congruent data set. c Theis recovery test (confined). When the pump is shut down after a pumping test, the water levels inside the pumping and observation wells will start to rise. This rise in water level is known as residual drawdown. Recovery-test measurements allow the transmissivity of the aquifer to be calculated, thereby providing an independent check on the results of the pumping test. Residual drawdown data can be more reliable than drawdown data because the recovery occurs at a constant rate, whereas constant discharge pumping is often difficult to achieve in the field. Residual drawdown data can be collected from both the pumping and observation wells.

Pe-0 1 Pe-01 Pe-02 Pe-02 Pe-03 Pe-03 Pe-03 Pe-03 Pe-04 Pe-04 Pe-04 Pe-05

14.12.2000 21.02.2001 14.12.200O 21.02.2001 14.12.2000 01.02.2001 07.02.2001 21.02.2001 15.02.2001 15.02.2002 21.02.2001 21.02.2001

8.1 7.1 7.5 7.1 7.3 6.9 7.0 6.8 7.0 7.0 7.4 7.3

114 59 35 20 10

22 18 14 10 6

106 171 100 62 51

84 39 33 20 20

46 39 23 19 14

342 300 206 99 88

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Figure 10.11. (a) Electric conductivity and (b) TDS measured over a horizontal cross-section of the experimental barrier at a depth of 1.5 m (mean values of the monitoring data for August 2 0 0 2 - M a r c h 2003 for monitoring wells PRB-1, 15, 16, 17, 18, 19, 20); location of the monitoring points see Figure 10.6.

Experimental iron barrier in P~cs, Hungary

269

groundwater pH when it enters zone 2 of the barrier, where iron is present in high concentration (see Figure 10. lOb). The electrical conductivity dropped from 1600 to 600 IxS/cm as the groundwater passed through the reactive zone (Fig. 10.1 l a). A similar behaviour was observed in the concentration of total dissolved solids (TDS) which was reduced by 700 mg/l of treated water, implying that 700 mg of precipitated compounds are formed per litre of treated groundwater (Fig. 10.11b). This effect is most likely a result of the precipitation of CaCO3 and MgCO3 caused by the increase in pH, and the removal of sulphate (Table 10.2). The calcium, magnesium and bicarbonate concentrations all decreased across the barrier (Fig. 10.12), proportionately by amounts that exceed the

(a)

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Figure 10.12. Calcium, magnesium and bicarbonate concentrations: (a) as measured in the observation period, and (b) as mean values over a horizontal cross-section of the experimental barrier at a depth of 1.5 m (mean values of the monitoring data for August 2002-March 2003 of monitoring wells PRB-1, 15, 16, 17, 18, 19, 20); location of the monitoring points see Figure 10.6.

2'/0

M. CsO'vdri et al.

drop in TDS. Sulphate is most probably transformed to sulphide in the reducing condition in the barrier. The estimated amount of precipitates that were formed within the barrier is rather high and could eventually lead to the loss of reactivity or hydraulic conductivity over the longer term. Therefore, the treatment of water with a high-alkalinity agent (e.g. high concentrations of dissolved carbonates) in iron-based PRBs is challenging. The redox potential, Eh, was measured in the laboratory, not immediately after sampling on-site. Therefore, the redox potential in the groundwater may have been slightly different. However, a tendency is apparent: extremely low values were measured in the two zones with reactive iron material, indicating that the redox potential was reducing within the experimental barrier (Fig. 10.13). Elevated iron concentrations up to 20 mg/l were detected in zone 1 of the PRB material, where the pH value was still comparatively low (see Table 10.2). At pH values above 9, the iron concentration is lower due to precipitation of iron minerals. The measured profile of iron along the cross-section of the experimental barrier and the dependency of the dissolved iron concentration on pH is shown in Figure 10.14. The reason that the PRB was constructed at this particular site in the Zsid valley of the Prcs region was that monitoring well Hb-Ol/l indicated a continuous increase of uranium contamination in the shallow groundwater. The monitoring well, located approximately 15 m downstream of the newly installed experimental PRB, was constructed well before the implementation of the PRB, and was therefore suitable for determining the effect of the experimental PRB on the local groundwater. Results from this monitoring well from before and after the implementation of the experimental PRB are compiled in Table 10.3 It can be clearly seen that the iron barrier affects the local groundwater composition significantly. The uranium concentration downstream of the iron barrier decreased by approximately 90%. The calcium, magnesium and bicarbonate concentrations also dropped significantly. The trends of the uranium concentrations measured in monitoring well Hb-Ol/1 are displayed in Figure 1O. 15. 3OO 25O

Zone 1 . . . . . . . . . . . . . . . . . .

20O 150

. . . . . . . . . . . . . . . . . .

0

I

J. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

r . . . . . . . . . . . . . . . . . /

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . . . . . . . . . . . . .

iI

i. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

~ . . . . . . . . . . . . . . . . . I I- . . . . . . . . . . . . . . . . .

I

I

I. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I

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-50

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

-100

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-150

. . . . . . . . . . . . . . . . . . .

I-

-200

r . . . . . . . . . . . . . . . . .

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. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

I

Zone 2

i. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

I

~" loo 50 LU

I

~

I

I 0

. . . . . . .

50

9

. . . . . . . .

9

i . . . . . . . . . . . . . . . . . 9

. . . . . . . .

I- . . . . . . . . . . . . . . . . .

9 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ,

100

150

21

Horizontal cross-section (cm) Figure 10.13. Redoxpotential Eh over a horizontal cross-section of the experimental barrier at a depth of

1.5 m (mean values of the monitoring data for August 2002-March 2003 of monitoring wells PRB-I, 15, 16, 17, 18, 19, 20); location of the monitoring points see Figure 10.6 (the last point is outside the reactive iron zone).

Experimental iron barrier in P~cs, Hungary (a)

25 E

2O

.o .t-.,

15

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l

Zone

~

11

i

--o--- ~. . . . . !. . . . . . . . . .

25

(b)

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E 20

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,,_.., c

.

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271

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.__o 15

.

r................

5

o

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U_

.

.

.

.

.

.

8

.

t.)

o

...........................................

10 ........................................... 5

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

I

0

I

,

50

100

6

150

200

O

n

7

|

i

i

8

9

10

11

pH

Horizontal cross-section (cm)

Figure 10.14. Concentration of dissolved iron: (a) over a horizontal cross-section of the experimental barrier at a depth of 1.5 m (mean values of the monitoring data for August 2002-March 2003 of monitoring wells PRB-I, 15, 16, 17, t8, 1,9, 20, location of the monitoring points see Figure 10.6) and (b) as a function of pH. The average results from all monitoring wells have been processed with the help of the contouring and surface mapping program SURFER for a graphical representation of the most probable isolines of the different measured components in a plan view and a vertical section through the barrier (Bakonyi and Ferenczi, 2003). Figures 10.16 and 10.17 give plan views of the distribution of HCO3, Ca, Mg, TDS, electric conductivity, pH and uranium in the vicinity of the barrier. The strong effect of the PRB on the water chemistry is clearly visible. Figure 10.18 shows the distribution of these quantities on a vertical section through the barrier. The form of the isolines indicates that the water flow in the barrier probably has components in both the horizontal and vertical directions resulting from the upward hydraulic gradient at the site.

Table 10.3. Concentration of selected groundwater constituents at monitoring well Hb-01/1, located approximately 15 m downstream of the experimental PRB, before and after implementation of the PRB system. Component

Unit

2000- 2002, average

After implementation of PRB a

Na K Ca Mg CI $04 C03 HCO3 TDS pH U Electrical conductivity

mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l Ixg/l ~S/cm

88 20 168 53 43 361 < 10 474 1030 7.1 718 1496

99 13.5 54 33 40 170 < 10 299 540 7.7 70 923

a

Averages over the monitoring period of August 2002-March 2003.

Change

-

11 6.5 114 20 3 19 l 175 490 0.6 648 573

M. Cs6vdri et al.

272 1200

1.8

1000

....

Imu

[..................................................................... O

O TDS

800 ......................................................... .......................................... O0 O Oo

600

400

....

~__o_____o_ . . . . . . . . .

O

.......

200

0

mm

01/96

OO

o_~ .......

l

,-

.,,--,.

01/97

. 01/98

9. . . . . . . . . .

O

O ._,=

1.5

o

O

_1__~ 9 .........

.e-m ......... o-tl ........................ 0 0 9 1 499 9

_e__m_. . . . . . . . . . 9 9 9

9. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

.

. 01/99

. 01/00

0.9 0 0

.............................................. .

,-mm

. 01/01

01/02

1.2 0.6

ijj

l:::l I--

0.3 0.0

01/03

Date Figure 10.15. Uranium concentration and TDS at monitoring well Hb-01/1 over a multi-year period. The start-up of the pilot barrier in August 2002 is clearly visible.

2. Hydraulic performance of the PRB The PRB has been built in Permian sandstone set in a rather bluff valley bottom. The valley bottom is filled with quaternary alluvial deposits. The approximately 3 - 5 m thick alluvial sediments were built up mainly from sandy and silty clay beds as a confined aquifer (Fig. 10.19) which exhibits an upward hydraulic gradient. Geo-electrical investigations showed that a highly resistive bedrock can be found below the approximately 5 - 7 m thick layer of sediments. The geological and geophysical investigation of the valley has led to the conclusion that a 3 - 4 m thick sandy layer with embedded clayey lenses can be found between an approximately 1-2.5 m thick clayey surface layer and the thick Permian sandstone bedrock. The upper bedrock is fractured. A simplified geological structure of the valley and its immediate vicinity, based on the site investigation, is shown in Figure 10.20. The hydraulic gradient is approximately 0.04-0.06 (average 0.05) and the groundwater flow velocity is about 0.086 m/d or 31 m/a. The hydraulic performance of the experimental PRB was measured using different methods. The results are listed in Table 10.4. The average value for the hydraulic conductivity kf is 3.36 x 10 -3 m/s. Instead of this value, the hydraulic conductivity of the surrounding subsoil must be used for the estimation of a realistic flow rate in the iron barrier, since this is the limiting factor for the flow rate through the barrier. Based on the site investigation described in Table 9.2, a hydraulic conductivity of k f - 2 x 10 -5 m/s can be assumed for the soil. With this value and an hydraulic gradient varying between 0.04 and 0.06, the following flow rates can be calculated for a flow-through area of 23.8 m 2 (6.8 m x 3.5 m)

Q - k f A i - 2 x 10 -5 X 23.8 X 0.04 -- 1.9 X 10 -5 m3/s = 600 m3/a Q - k f A i - 2 x 10 -5 x 23.8 x 0.06 = 2.9 x 10 -5 m3/s = 900 m3/a

Experimental iron barrier in Pdcs, Hungary

Figure !0.16. Isolines of bicarbonate, calcium, magnesium and TDS in and around the PRB.

273

274

Figure 10.16. (continued).

M. Cs6vdri et al.


275

Figure 10.17. Isolines of pH, specific electric conductivity and uranium concentration in and around the PRB.

276

M. Csr"vrri et al.

Figure 10.17. (continued).

It should be noted that the accuracy of such an estimate is limited by a number of difficult-to-assess factors, such as uneven distribution of flow due to inhomogeneous layering of the soil, and seasonal changes in the site hydrology. The hydraulic situation on the test site after the installation of the PRB is shown in Figure 10.21. There was a small water flow on the surface around the four monitoring wells PRB 2 1 - 2 4 during the measurements as a result of the upward hydraulic gradient. Based on the flow rate and the reduction in TDS in the barrier, the effect of the precipitation on the hydraulic properties of the experimental PRB can be assessed quantitatively. Assuming a continuous flow rate of 750 m3/year (the arithmetic mean of 600 and 900 m3/a calculated above) and a reduction of TDS of 700 mg/l by the elemental iron, 525 kg of solids will precipitate within the iron barrier each year. Using values of 2.75 g/cm 3 as an average density of the precipitates and 30% for the porosity of the iron/sand mixtures, the annual loss in porosity in the experimental iron barrier can be calculated as approximately 1.6% which translates to 100/1.6 -- 62 years of operational life. This is a bulk average value since it assumes even distribution of the precipitates through the barrier and will certainly underestimate the amount of precipitation at critical locations in the barrier. Precipitates will form first at the upstream interface ("entry front"), but in time the "precipitation front" will move deeper into the PRB. The amount of precipitation in unit PRB volume will therefore be the maximum at the entry front and will decrease

Experimental iron barrier in Pgcs, Hungary

Figure 10.18. Verticalcross-sections of the distribution of selected groundwater constituents in the PRB. until it reaches zero at the precipitation front. A PRB failure due to precipitation will occur when its permeability decreases to that of the surrounding soil. The distance that the precipitation front has travelled through the PRB by this time defines the maximum functional thickness of the barrier. Larger thicknesses are not justified, they

278 M. Cs6vdri et al.

r~

~,-~ r~

0 0 ~,-~

r~

!

r~

0

~,..~ em 9 0 O

r13

Experimental iron barrier in Pdcs, Hungary

279

Figure 10.20. Simplifiedgeological structure of the Zsid valley with hydraulic parameters.

may only be necessary when the dimensions of the plant used for excavation make them unavoidable. Experience of other researchers (Yabusaki et al., 2001) substantiate our findings: the bulk of precipitates form at or near the upstream interface where the groundwater enters the reactive i~on material. Therefore, by concentrating the precipitates near the upstream interface, the actual effect on the hydraulic properties of the experimental barrier system will be more pronounced and have a more rapid influence on the overall performance of the system. The operational life estimate also ignores the possibility that the barrier reactivity could be lost due to coating of the iron particles before the porosity is completely blocked. Future monitoring and - if possible sampling of iron material from the installed barrier will provide further data on the long-term behaviour of the system.

Table 10.4. Results of different measurement of the hydraulic conductivity kf for the experimental PRB. Measurement method

kf (m/s)

Theis Recovery Test Cooper-Jacob Steptest Theis Steptest Cooper-Jacob Distance-Drawdown Neuman Method Cooper-Jacob Time-Drawdown

3.98 x 2.01 x 1.95 x 5.43 x 1.11 x 5.67 x

10 -3 10 -3 10 -3 10 -3 10 -3 10 -3

280

M. CsO'vdri et al.

Figure 10.21. Hydraulic situation at the site after installation of the experimental iron barrier:

groundwater level and flow direction around the PRB (15.04.2003).

E. Conclusions The operation of the PRB at the site near P6cs has shown promising results in the first few months after installation. Monitoring data indicate that the experimental PRB operates very efficiently regarding the removal of U from the local groundwater.


281

Following the construction of the experimental PRB, the uranium concentration in nearby monitoring wells has dropped significantly from values around 1000 txg/l to values down below 100 txg/l. Within the barrier material itself uranium concentration has decreased to less than 10 Ixg/l. Further monitoring is needed to evaluate long-term performance, especially with regard to the effect of mineral precipitation. An operational period of the test barrier of 168 years can be estimated if only the amount of iron and the volume of treated groundwater is taken into account (38 t or 680 x 103 mol iron, stoichiometric factor U:Fe -- 1:1390, see Chapter 3, 700 m3/year water with 1 mg or 4.2 x 10 -6 mol uranium/l). The change in water chemistry regarding TDS indicates that 700 mg of carbonates per litre of treated water are precipitated as a result of pH value increase. It is therefore most likely that changes in the hydraulic conditions of the system, rather than the amount of reactive material in the barrier, will be the limiting factors on the barrier's operational life. A loss in total porosity of 1.5% per year was calculated for the experimental iron barrier. This would translate to 62 years of operational life assuming a homogeneous distribution of precipitates, but the operational life will almost certainly be shorter because precipitation occurs preferentially at the upstream side of the barrier and coating of the reactive particles reduces their reactivity. The most important achievement is, however, that the installation of the PRB has fundamentally improved the situation regarding the main contaminant uranium. Therefore a PRB can be recommended on basis of the data collected, even after allowing for the short period of operation, as a significant attenuation in the uranium concentration in groundwater was observed when otherwise it would have increased.

References Bakonyi, ,/~., Ferenczi, B., 2003. Treatment of uranium-contaminated groundwater using reactive barriers. Diploma Thesis, Technical Department, Pollack Mih~ily University of P6cs, Hungary. Cantrell, K.J., Kaplan, D.I., Wietsma, T.W., 1995. Zero-valent iron for the in-situ remediation of selected metals in groundwater. J. Hazard. Mater. 42, 201-212. Fiedor, J.N., Bostick, W.D., Jarabek, R.J., Farrell, J., 1998. Understanding the mechanism of uranium removal from groundwater by zero-valent iron using X-ray photoelectron spectroscopy. Environ. Sci. Technol. 32, 1466-1473. Merkel, B., Sperling, B., 1998. Hydrogeochemische Stoffsysteme II. Schriftenreihe des Deutschen Verbandes ftir Wasserwirtschaft und Kulturbau e.V. (DVWK), 117, Bonn. Morrison, S.J., 2003. Performance evaluation of a permeable reactive barrier using reaction products as tracers. Environ. Sci. Technol. 37, 2302-2309. Morrison, S.J., Metzler, D.R., Carpenter, C.E., 2001. Uranium precipitation in a permeable reactive barrier by progressive irreversible dissolution of zero-valent iron. Environ. Sci. Technol. 35, 385-390. Naftz, D.L., Morrison, S.J., Davis, J.A., Fuller, C.C., 2002. Handbook of Groundwater Remediation using Permeable Reactive Barriers. Elsevier Science, Amsterdam, p. 539. Phillips, D.H., Gu, B., Watson, D.B., Roh, Y., Liang, L., Lee, S.Y., 2000. Performance evaluation of a zerovalent iron reactive barrier: mineralogical characteristics. Environ. Sci. Technol. 34, 4169-4176. Yabusaki, S., Cantrell, K., Sass, B., Steefel, C., 2001. Multicomponent reactive transport in an in situ zerovalent iron cell. Environ. Sci. Technol. 35, 1493-1503.


K.E. Roehl, T. Meggyes,F.-G. Simon, D.I. Stewart, editors 9 2005 Elsevier B.V. All rights reserved.

283

Chapter 11 Installation and operation of an Adsorptive Reactor and Barrier (AR&B) system in Brunn am Gebirge, Austria Peter Niederbacher and Manfred Nahold

A. Introduction

Industrial land use since the early stages of industrialisation has, in many cases, caused environmental impacts on the soil and groundwater. Nowadays these industrial brownfield sites, which are often situated in well-developed urban areas, are being reused for new building projects. This process has been called "brownfield recycling". At the former site of a tar factory and linoleum production plant in Brunn am Gebirge near Vienna, Austria, soil and groundwater contaminations caused mainly by residuals of tar production were detected during the starting phase of construction for a business park. To allow the proposed land use without undue delay, a remediation scheme that included an innovative groundwater treatment and protection system was adopted. The system design is based on the permeable reactive barrier (PRB) concept. It consists of a hydraulic barrier with four openings, each connected to an adsorptive reactor filled with activated carbon (AC). The flow through the system is driven by the hydraulic gradient between the contaminated (upstream) area and the runoff of the system (downstream) to an artificial groundwater pond, which is connected to the river system. The technical facilities are fully integrated into the business park, which was built at the location. The system implementation was financed by a private investor and supported by funds of the Republic of Austria. The start of the European research project PEREBAR ("Long-Term Performance of Permeable Reactive Barriers Used for the Remediation of Contaminated Groundwater", see Preface; Niederbacher, 2000 and 2001) in 1999 coincided with the start-up of the Adsorptive Reactor and Barrier (AR&B) system in Brunn am Gebirge. This was an excellent opportunity to choose the site for closer investigation during the start-up phase and first years of operation of a full-scale PRB system. The investigation was undertaken in cooperation with the consultant group Niederbacher/Mapag who is responsible for system design and monitoring. The monitoring facilities installed in the AR&B system allowed the operation of a project scale PRB to be investigated under real world conditions. Besides routine monitoring for the control of hydraulic function and purification efficacy, the investigations performed in the course of the PEREBAR project allowed a detailed study of general conditions which are influencing the long-term behaviour of the AR&B system.

284

P. Niederbacher and M. Nahold

B. General description At the former site of a tar factory and linoleum production plant near Vienna, Austria, a groundwater protection and remediation system based on the PRB concept was installed and put into operation in October 1999. The groundwater in the area of the former industrial plant showed significant contamination with polycyclic aromatic hydrocarbons (PAH), hydrocarbons (HC), BTEX, phenols, and - to a minor extent - chlorinated hydrocarbons (CHC). The system is called after the location "Adsorptive Reactor and Barrier (AR&B) system Brunn am Gebirge". The main component of the system is a hydraulic barrier, which is combined with four in situ adsorptive reactors containing activated carbon. The runoff from the system is released to an artificial groundwater lake with an outlet to the river system. l. Location and case history

The site of the groundwater remediation system is situated close to the southern city limits of Vienna, in Brunn am Gebirge, Austria, south of highway A21 (Fig. 11.1). The site, which covers an area of 60,000 m 2, was the location of a former tar plant which operated from 1878 to 1932. The tar plant was replaced by a linoleum production plant, which existed until the 1960s. Buildings and installations were dismantled in 1974. The mix of contaminants is related to the industrial activities as well as to incidents such as spills during almost 100 years of industrial operation. The contamination of the unsaturated and saturated zones was discovered in 1997 during excavation for services that formed part of the infrastructure of the business park. In the environmental investigation that followed (1997-1998), 23 wells for groundwater sampling (depths 6 - 1 8 m) and more than 50 local excavations (depth 2 - 5 m) were made. Selected investigation pits, which reached below the groundwater table, were equipped with filter hoses for groundwater observation. Based on these data a two-step program clean-up of the unsaturated zone and implementation of the groundwater protection and remediation system - was planned by the consultant team Niederbacher/Mapag and co-ordinated with the landowner and appropriate authorities. After excavation of

Figure 11.1. Locationof Brunn am Gebirge, Austria.

Installation and operation of an Adsorptive Reactor and Barrier (AR&B) system

285

the hazardous tar waste from the unsaturated zone (August 1998-June 1999) and installation of the PRB system (February 1999-August 1999) the site was integrated into a business park owned by a private investor.

2. Geological-hydrogeological set-up The site is located near the western margin of the southern Vienna Basin, an Alpine pullapart structure, filled with Tertiary sediments. At the margin of the Vienna Basin, normal faults like the "Sollenau Fault" are known from drilling and seismic data. This mentioned tectonic structure occurs in the western vicinity of the site. The general profile shows anthropogenic deposits on top ( 0 - 2 m) and inhomogeneous alluvial sediments (sandy, silty gravel) with a thickness of 3 - 6 m. These clastic sediments are underlain by shales of mid-Pannonian age with intercalation of coarser layers in which artesian water occurs. The relief of the top of the Tertiary age sediments is a roof-like dipping from west to east, with a north-south oriented depression, and a rising shoulder towards the east where the confining layer reaches the surface. This relief can be explained by substructures of the nearby "Sollenau Fault" which is orientated in the same direction. These older sediments were partially eroded and covered by Quaternary age deposits. The groundwater table is 2 - 4 m below ground level. The base of the aquifer is at a depth of 3 - 6 m. Pumping tests showed permeabilities of kf - 3 • 10 - 3 - l • 10 -5 m/s in the quaternary sediments. The initial natural groundwater flow was oriented from west to east, bending towards southeast following the depression of the relief of the underlying layer, which is acting as an aquitard (Fig. 11.2). This caused a migration of contaminants with the groundwater flow towards southeast.

3. Contamination of the unsaturated zone The results of the environmental investigation of the site showed anthropogenic deposits of solid waste material from tar production, similar to those often found at manufactured

Figure 11.2. Site-specificset-up: relief of the underlying aquitard, initial status of groundwater flow.

286

P. Niederbacher and M. Nahold

gas plant sites, over an area of approximately 30,000 m 2. Residues from the tar refinery included: tar, solid naphthalene, ash and clinker, treatment sludge, impregnation of soil with tar oil, phenols, aromatic hydrocarbons and hydrocarbon (spills) of middle and heavy oils, whereas residues from the linoleum production included: waste linoleum, linoleum raw material (cork, linseed oil and mineral pigments containing heavy metals) and hydrocarbons (from spills). The concentrations of the contaminants exceeded the action threshold values stated in the environmental regulations (e.g. ONORM S 2088), for the relevant parameters such as PAHs, HCs, phenols and heavy metals, over a large area. The analytical results showed maximum total petroleum hydrocarbon (TPH) contents of tens of thousands of mg/kg and PAH contents as high as 8000mg/kg. The solid waste material was excavated, separated on-site and then sent for treatment or disposal according to the relevant regulations. A total of approximately 80,000 ton of waste and residues were excavated. 4. Groundwater contamination Related to the waste deposits and spills from the industrial use of the site was a plume of contaminated groundwater (Table 11.1) which was drifting with the natural groundwater flow towards south/southeast. Groundwater investigations indicated that there was significant contamination of the saturated zone (groundwater fluctuation zone) over an area of more than 35,000 m 2. A containment solution would have led to restrictions on the intended use of the area as building land. The treatment of the contaminated groundwater was not realistic due to the level of groundwater fluctuation. The financing of an extensive excavation and removal of contaminants in both the unsaturated and saturated zones was substantially beyond the capabilities of the private investor. As a result of decades of in situ alteration (biodegradation) the hydrocarbons occurred as widely degraded immobile residuals in the unsaturated and saturated zones. Therefore, no free phase was observed in the wells and investigation pits. The source of the CHC input into the groundwater (probably caused by manipulations during the use as linoleum production plant) could not be located. An input through the sewer system or sinks seems to be possible. The migration and current distribution of the groundwater contamination was driven by the initial flow direction of the groundwater from W ~ E, bending to NW ~ SE (see Fig. 11.2). The flow follows the erosional depression on the top of the Tertiary sediments (underlying aquitard). The initial runoff of the shallow groundwater horizon was towards a streamlet south of the location (Fig. 11.3).

C. Site assessment

l. Environmental and hydrogeological investigations The environmental site assessment was based on the geotechnical, hydrogeological and environmental-chemical data from the investigation phases in 1997-1998. Data were

Table 11.1. Groundwater contamination at selected locations (investigation wells B9805 and B7 in the centre of the pollution source, B9806 and B9808 at the edge of the plume downstream of the pollution centre and B9803 upstream of the pollution centre) ~~

Locationlwell

ETPH

Benzene

Xylenes

--

mn)

(Pm

(Pm

3.3 6.8 0911998

19 29 0611998

435 640 0611998

35 50 0511989

0611998

0.6 3 1111997

1.15 4 0511998

36 62 0411999

0.5 5 Ill1997

0.02 0.17 0512001

0.14 3 0911998

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