long-term performance of permeable reactive barriers, volume 7 (trace metals and other contaminants...

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

2 0 0 5

E L S E V I E R

A m s t e r d a m - B o s t o n - H e i d e l b e r g - L o n d o n - N e w Y o r k - O x f o r d

P a r i s - S a n D i e g o - S a n F r a n c i s c o - S i n g a p o r e - S y d n e y - T o k y o

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Contents

Preface Contributors About the editors Acknowledgements

1. 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

xi XV

xvii xix

1 1 1

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

vi Contents

o

,

D

4 Hydrated lime 5 Elemental iron 6 Iron oxides 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

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

On-site Column Experiments A Introduction B Columns in monitoring wells

1 Experimental set-up 2 Results

67 68 71 72 73 76

77 77 81 81 83 86 87 90 99

100 104

111 111 111 111 114 114 115 116 116 119

121 124 125 127 129 130 131 132

137 137 138 138 138

Contents vii

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

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

.

10.

11.

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

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

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

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

227 227 229 229 232 242 242 242 243 253 257

261 261 261 262 267 267 272 280 281

283 283 284 284 285 285 286

Contents ix

C 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

296 299 299 299 301

302 303 307 308

12. 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

315 316 316 318

319 320 320

Index 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 full- scale 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 non- destructive 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 uranium- contaminated 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, cost- effective 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)

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

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

J6zsef B6hm Department of Process Engineering, Miskolc University, Egyetemvfiros, H-3515 Miskolc, Hungary

David E. Bryant School of Chemistry, University of Leeds, Woodhouse Lane, Leeds, LS2 9JT, United Kingdom

J6zsef Csicsdk MECSEKI~RC Rt., Eszterg~ir Lajos u. 19, H-7633 P6cs, Hungary

Mihdly Cs6vdri MECSEK-0KO Rt., Eszterg~ir Lajos u. 19, H-7633 P6cs, Hungary

Kurt Czurda Department of Applied Geology (AGK), Karlsruhe University, Kaiser- strasse 12, D-76128 Karlsruhe, Germany

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

Hansj6rg Fader Germany

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

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

Imre Gombk6t6 Department of Process Engineering, Miskolc University, Egyetem- v~ros, H-3515 Miskolc, Hungary

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

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

xvi Contributors

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

Stefan Ludwig Germany

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

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

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

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

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

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

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

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

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

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

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

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

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 Manage- ment 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

Permeable 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. Schematic depiction 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:

�9 Degradation. 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.

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 "funnel- and-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 non- pumping 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.


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 CrO 2- 8H + Fe3+ Cr3+ Fe~solid) + + ~ + + 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

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

(1.5)

(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 ~ +0.851V; 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).


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 co- precipitation. 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)

Pd Of(c) R = 1 -~ (1.9)

n 0c

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 10OO m2/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 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.

�9 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.

�9 Availabil i ty 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.

�9 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.

�9 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.

�9 Safety. Handling of the material should not generate any significant risk to workers' health.


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 palladium- modified 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 co- solutes 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


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


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 m2• 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 full- scale 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 up- gradient 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 lOOO mg/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 up- gradient 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 -10 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


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 >99% (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 pre- supposition for a suitable reactive tracer is the specific reactivity solely with iron. A good


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


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


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


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 0-66 to 690-2300 mg/l. The PRB converted the aquifer from acid-producing to acid-consuming. The populations of sulphate-reducing bacteria were lO,OOO times 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)


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 5-16 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


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).

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Grathwohl, P., Peschik, G., 1997. Permeable sorptive walls for treatment of hydrophobic organic contaminant plumes in groundwater, Proceedings of the International Containment Technology Conference, February 1997, St. Petersburg, Florida, USA, pp. 711-717.

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24 K.E. Roeh l et al.

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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. 27

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 continuous reactive barriers, �9 funnel-and-gate systems, �9 arrays of wells, �9 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 (two- phase 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).

28 T. Meggyes

�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 self- hardening 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 36-48 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). ~

Soil Material Depth (m) Thickness Permeability Experience (m) t d S )

Principle cut-off wall system

Excavation of soil and Single-phase placement of dlaphrdgm Wall sealing material

Two-phase diaphragm wall

Limited suitability for peat and humic acids



No limitation if casing is applied

Bentonite-cement mix with or without filler

Bentonite slurry, soilcrete

Approximately 35

0.4- 1.5

0.4- 1.5

> 0.6

0.6-0.8

0.05-0.20

0.01 -0.02

>0.4

adjustable

>0.8

>0.8-1 .o

5 1 x 1 0 I ( ' Landfill 28 years

<5x 10-10 > 50 Landfill 20 years

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

< 5 x lo-'"

Composite diaphragm wall

Bentonite- cement mix, dditional asealers

Soilcrete, concrete

Approximately 30

Landfill 20 years

Interlocking bored-pile diaphragm wall

Displacement Thin wall of soil and installation of sealing material

Sheet-pile wall

Approximately 20

Landfill 14 years

I 1 x lop9 Landfill 18 years Bentonite- cement mix with fillers

18-23

Suitable for pile dnving

Steel 15-20

15-20

Contaminated

Contaminated land 18 years

land 19 years

Hydraulic and foundation engineering

Hydraulic and foundation engineering

Civil and mining engineering

-

5 1 x 10-9 Driven cut-of wall

Reducing Injection wall permeability of soil in place

Jet grouting, HPI

Soilcrete, concrete

5 1 x 1 0 Injectable Cement, achy -cement mix, silicate gels

Bentonite-cement amix with and without filler

Liquid nitrogen

> 100

5 1 x 10-l0 Very fine-grained > 100

Frozen wall > 100

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, cement- bentonite 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 70- 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 50-70 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 50-180 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 clay- cement-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 60-80 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 cut- off 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,


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; �9 single-phase slurries; �9 two-phase slurries.

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 high- solids content. A well-proven mixture for thin wall slurries is (Arz, 1988)

�9 25 kg bentonite �9 175 kg Portland cement �9 800 kg rock flour �9 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:

�9 42 kg Na-bentonite �9 200 kg kiln cement �9 917 1 water

Calcium bentonites are also used to produce cut-off wall slurries with high solids content. A well-proven mixture is

�9 1 6 5 kg Ca-bentonite �9 144 kg kiln cement �9 826 1 water

Ca-bentonites exhibit lower swelling and provide more stable suspensions than Na- bentonites. 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 cut- off 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 cut- off 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):


�9 replaceability of the reactive materials; �9 higher permeability than the surrounding reservoir; �9 stability against fines washed into barriers from the surrounding soil; �9 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.


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

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

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

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

38 T. Meggyes

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 earth- moving 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.


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

Contamination CHC GAC loading GAC volume Loading time

Investment Depreciation time Depreciation costs Interest rate Interest costs

Total capital costs

Electricity (l 0 W/h) Operation (manpower) GAC Repair- maintenance Discharge of groundwater (0.50 C/m 3)

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

funnel length: 270 m

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

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

100,000 C/year

m

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

Monitoring

Total operational costs

Total costs Specific costs per treated

volume (175,200 m3/year)

10,000 C/year

60,000 C/year

160 000 C/year 0.91 C/m 3

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


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

Slurry pump I[ ' , J l ",l !

, illli - - ,

I ,, _ - - 5 ' I - -

\ Stop-end tube

Jet cutting head 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).


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 funnel- and-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 non- continuous 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 Materials, equipment and contractors are easily available. �9 Well installation is straightforward. �9 Greater depths (> 50 m) are feasible. �9 Only a fraction of the reactive material is needed in comparison to continuous walls. �9 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 50- 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 Maximum open surface area of the DART housing. �9 Maximum groundwater flow rate through the reactive material. �9 Easy production of DARTs with various diameters and lengths. �9 Easy placement of the DART. �9 Easy access to the reactive material. �9 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

�9 Injection into existing pores in the ground. �9 Ground fracturing.

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


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 down- hole 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. Schematic biobarrier configuration (Hiebert, 1998).

Preliminary economic estimates by Hiebert (1998) indicated a cost range of US$ 6.5- 10.7 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.


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 MTBE- degrading 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.

48 T. Meggyes

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) Costs a

Caisson-based construction > 15 Mandrel-based construction 12-15 Continuous trenching 8 Jet technology 60 Deep soil mixing 50 Hydraulic fracturing 25-40 Vibrating beam 30

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/m 2

a Without mobilisation costs, based on 1C = 1.255.


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 Sta te-of- the-Art report "Landfil l Liner Systems"; Land Con tamina t ion and Reclamation for permission to use parts of "Removal 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 Remediat ion".

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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 Hydroxyapatite Activated carbon Ca(OH)2 Elemental zeolitic tuff iron

Particle size (mm) a 0-1.2 0-1.25 1.5-4 0-0.59 0.35-1.2 Density (kg/dm3) a 2.16 0.5 0.48 NA 6.7-7.2 Specific surface 15.6 65 1000 NA 0.048

area (m2/g) a Paste pH b 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 Si 4+ 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

o

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.


(a) 100

~, 80- v m

> o 60- E

E 40- t - -

l,...

D 20-

0 0

41'

i i

4 6 pH

! !

8 10 12

(b) 100

90- v m

o 80- E

E 70- .m

C t~

:D 60-

50 i I i

O 1 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 -OH) and aluminol (AI -OH) 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) (3.1)

SOH ~ + O H - ,---, SO- + H20 (in alkaline solutions) (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.


100%

80% v

tO m 60%

E E 40% t ~ L

E) 20%

0%

: U 0 7 + ~ " ~ (UO2)3(OH) ~

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

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

0

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

-I .... i-I ..... C~ 3(aq)

2 4 6 8 10 12 pH

Figure 3.2. Uranium speciation 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 t -

(l.)

o - 30 13 .

"~ -35 N

-40

-45 o i

� 9 1 4 9

�9 _�9

- i

I I i

i i i i i

4 6 8 10 12 14 pH (final)

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


1,5

1,0- (.9 v

0 , 5 - 03

0 , 0 - t - O (D o -0,5 -

= - 1 , 0 -~

-1,5 -

-2,0 -

-2,5 4

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

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 uranium- beating 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 uranium- beating 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 carbonato- complex 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.


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


v

,m O

{3.

E .=2

t - -

t ~

1 0 0 % "

8 0 %

6 0 % -

4 0 % -

2 0 % -

0%, 2

�9 / U O 2 ( C O 3 ) 3 -

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

UOzCO3(aq) .......................

uo2(co3)~-

....... i . . . . . . . . . . .. i ................ ' ~ . v . . . - . , . i- - . . ..i.. . . - . ~

4 6 8 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


(a) Effect of pH variation (contact time 2 h) 100

95 O E 90 E .=_ 85- t--

l,,.. D 80

2

4~

; 8 1'0 pH

(b) Effect of contact time (pH 7.6) 100

> 95 O

~ 9o E ._~ 85

D 80

__o.-

1'0 9'0 30 Time (h)

(c) Effect of the solid/liquid ratio (pH 8.5, contact time 1 h)

~lOO �9

95 > O E �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 1OOO txg/l of uranium (VI).

1E+00

1 E-01

o 1E-02 E " 1E-03 O

t~

1E-04 o

"- 1E-05 O o

1 E-O6

1 E-07

. . . . : : ~ -

~ ~ m e t

" ' ~ i B m .A_ . . . .

" " " " ~ ~ ~ "~, : - O - - "~- . . . . . .

_ .... ~ " .... . O O

_ " ~ _ o

3 4 ; ; ; 8 ; 10 pH

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-) .


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 P04 surface species:

\ - - C a

\ C a O H ~ / " 0 __ \p~.O

/ \ o " \ - - C a ~ O- /

In aqueous solut ions the a c i d - b a s e react ions that are responsib le fo r the surface properties of HAP are as follows

= P O - + H + ,--, =POH ~ with log K = 6.68 (3.4)

- C a _ O H + ~ - C a _ O H ~ + H + with log K -- -9 .66 (3.5)

The surface speciation diagram of HAP in aqueous solutions is presented in Figure 3.9. Negatively charged PO- sites and neutral Ca -OH ~ sites predominate in alkaline solutions (Fig. 3.9). Therefore, the surface charge of HAP in alkaline solutions is negative. On the

1 0 0 %

80% -

"- 60% - 0 O

u_ 40% -

20% -

0% - 4

- - - CaOH~

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)).


30

2O

v

�9 - 0 C (1)

o -10 Q.

- 2 0 U

-30

-40 I I I I I 5 6 7 8 9 10 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%

80%

if)

�9 - 60% O c~

E 40% .2 C

:D 20%

0%

. . . . . . . . . . . . . . . . . . . . . . u%(c%)]-

_

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

. . . . . . . . . . _ , / _ . . . . .

~ ~ ' UO2H2PO 4 ' ~

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

, / uo~Po~

2 3 4 5 6 7 8 9 pH

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


8

~ 6 0

E 4 v

~ 2

-~ 0 0 ffl

E -2 "E -4

~ -6 o - 8

-10 i i I I

2 4 6 8 10 12 pH

---O-- Precipitation as uranyl phosphate

- - I - - Precipitation as autunite

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 10-3M total phosphates and 1.6 x 10-3M 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 (Ksp- 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 10-3M 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


10

~" 8- O E 6-

4- . Q _= O (n 2- E 0- . m (3

r j - 2 - E ~ o - 4 -

-6 ~l t I I i 2 6 8 10 12

pH

--<>-- Calcium phosphate

- - i - - Calcium carbonate

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 CaUO4 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


(a) ,40 35

"O 30 > "5 25 oo ._~ 20

"O E 15 ._ 10 t -

5 L

D O

�9 �9 �9 _ k . . . . . . .

�9 pH 2.5 [] pH 7 �9 pH 13

. I - t n [ ] . ~ ~ , [ ] ~ 13 , ,~1 ~ 0

0 200 400 600 800 100O Time (h)

(b) 350 300 =,

v 250 . . . . �9 �9 _A O

200 ~ - - i . pH 2.5 ~ . . . . . . . . . . . [] pH7 E .-= 100 �9 pH 13 = - - - limit 5O

O ;1 [] 0 200 400 600 800 1000

Time (h)

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.


( a ) ~ 2 0

> 1 5 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . O

E .~ 10 . . . . . . . - * . . . . . . . . . . . . . . . . . . . . . . . . .

._= 5 r

D O o 2 4 6 8 lo 12 o + 10

p H A C c o n c e n t r a t i o n ( g / I )

(b) o~" 30 v

--~ 2 5 . . . . . . . . . . . . . -~ . . . . . . . . . . . . . . . . . . .

S 2 0 . . . . . . . . . . . . . . . . . . . . . . . . . . . . -*- . . . . E 1 5 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . t _

E 1 O . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

E 5 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

5 o 12

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


Table 3.2. Uranium concentration in solution calculated using the geochemical programme PHREEQC.

Uranium species Concentration (mol/l)

UO2OH + 2.29 x 10-13 (UO2)3(OH) + 6.40 • 10 -14

U02(C03) 2- 1.27 • 10 -16

U02C03 3.23 X 10-17 U02(C03) 4- 6.91 x 10 -18 U022+ 2.44 • 10 -20

(UO2)z(OH) 2+ 4.63 X 10 -21

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 uranium- loaded 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.


(a) Effect of pH variation (contact time 2 h) ,oo _ ~ ~ ; ~ :

-~ 80 . . . . . . . . . . . . . . . . . . O E 60 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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

"E 20 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . t~

~ o

pH

(b) Effect of contact time (pH 7) ~ 1 0 0

, 80 4o 6o ! ii ! ! ii ii ii ! ! ii ii ii ii ii ii ii ii ii ii ii ! ! ! ! E 20

:::9 0 . . . . 0 400 800 1200 1600

Time (h)

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

~ioo --_ .-- -- ~ -~-- . . . . I , - . . . .

> 95 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . O E

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

D 80 0 1 2 3 4 5

Solid/liquid ratio (g/I)

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

~o " 6080 i i i i i~iiiii i '- 40 ._~:~ ~o ~ iiiii~iiiiii

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

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 uranium attenuation mechanisms by e lementa l iron, the iron

corrosion processes which are inherent to e lemental iron in aqueous envi ronment have to

be considered. Iron corrosion depends largely on the condit ions prevai l ing in the aqueous

medium and is initiated as shown in the fol lowing 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)

( a ) ~ l 0 "O

r 8 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

.a o 6 �9 �9 �9 �9 E

4 E 2 ~ . . . . o pH 5.5_ I �9 -,._ _ ~ . . . . . . . . . . . . . . . . . pH13.41

0 D 0 200 400 600 800 1000

Time (h)

(b) ~ 1 0 0

.8ot.o: 6o .............. - .... " .... : ..... I O

E 40-;-- . . . . . . . . . . . . . . . . . . . . o pa 5 5 - I ._.a ~ . . . . . . . . pH 13.4_1 = 20

| _ _ limit l Z) O -P - -o ~ , �9 , , , I

0 200 400 600 800 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).


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 +

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

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

(3.8)

(3.9)

(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 non- reductive 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


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 -OH 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- UO2 2+ *-+ (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


7

5 �9 �9

. "* . [ ' . . , -

o olL.' .** �9

-1

-2 2 & & ; 1'o 11

pH

Figure 3.18. Distribution coefficient 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


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 zero- valent 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/.


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.


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.


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 (pump- and-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

> g U . J

1000 800- 600- 400- 200-

0 - -200 - -400 -600

0

U(lV) .................... Fe/Fe 2+ P/-12 ; ~ C 0 2 i " ..............................

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 dissolution- precipitation mechanism. If autunite or chernikovite is the precipitation product the reaction sequence would be:

Cas(PO4)3OH--~ 5Ca 2+ + 3P03- + OH- (4.5)


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 - - - H 2 ( U O 2 ) 2 ( P O 4 ) 2 (4.8)

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)

~ -03P-OH + + 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.


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. Sketch and photograph of the test columns used in experiments on uranium removal from artificial groundwater.


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.


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) p~g/l 490 385 Na + mg/l 63.7 44 K + mg/l 16.0 14 Ca 2+ mg/l 134 140 Mg 2+ mg/l 39.4 40 CI- mg/l 47.5 32 SO 2- mg/l 234 253 CO 2- mg/l 11.4 10 HCO3 mg/l 443 412 Carbonate hardness ~ 21 18.9 Total hardness ~ 28 28.8 pH - 6.8 6.8 Electrical conductivity IxS/cm 909 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 mm 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 0 . 8 v

g -.~ 0.6 ..Q

0 .4

0 .2

a .

0.0

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

!!!!!!!!! ' ' ' ' I ' ' ' ' I ' ' ' ' I ' ' ' '

O 5 10 15 20 Time (days)

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


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 10-5-10 -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 HAP 20/80 306 C3 HAP 10/90 158 C4 Fe ~ 70/30 1795 C5 Fe ~ 50/50 1160 C6 Fe ~ 30/70 587


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).


40

35

30

g 25

~ 2o

E

l

5 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

" Photometry

- - ~ - I C P - M S

�9 ICP-AES

10

5

0 i i i i i i

0 2000 4000 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


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 long- term 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

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

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

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


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


Discharge

Test column filled with sand and iron

Uranium solution reservoir

I Pulse counter I

d collimator

Sodium iodide ~,-detector

Figure 4.9. Schematic representation 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

10 4

0 3 0

E 10 2 z

101

. . . . I . . . . l . . . . i . . . . i ~'----V-,,,~ 237 u

j ~ l l l j , , ~ , , , , ~ 3 J J t I - 132Te 1321

500 1000 1500 Energy (keY)

Figure 4.10. Photon emission spectrum of the test solution.


Table 4.5. Radionuclides produced through photoactivation of natural uranium.

Nuclide Half-life Activity (Bq/l) Count rate (s-l) Partition of the total count rate (%)

237U 6.75 d 125,914 1320 95.16 93y 10.1 h 128 0.9 0.06 95Zr 64 d 661 1.2 0.09 97Zr 16.8 h 336 1.0 0.07 99M0 66 h 3247 1.4 0.10 99mTc 66 h 3113 5.4 0.39 lO3Ru 39.3 d 999 4.0 0.29 127Sb 3.85 d 81 0.5 0.04 131Te 30 h 1000 9.9 0.71 132Te 76.3 h 3161 12.3 0.89 131I 8.02 d 2245 5.1 0.37 132I 76 h 2313 9.3 0.67 133I 21 h 831 1.1 0.08 140Ba 12.8 d 2774 2.2 0.16 140La 40.3 h 2488 1.8 0.13 141Ce 32.5 d 1460 7.3 0.53 143Ce 33 h 1497 1.8 0.13 153Sm 46.75 h 1141 2.0 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 237U without 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 237U concentrations 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


250000

200000 / ~

E �9 g.~_ 15oooo ~~'N~\ a~ 100000 [ ~

13_

50000

0 j ,

0 2

1: 96h , 2 : 1 2 0 h

3: 168h \ 4:216 h

5:336 h \ \ 6:480 h ~, \ 7:642 h

, i , '

4 6 8 10 Position (cm)

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 maximum concentrations at a distance of between 20 and 40 mm from the inlet.

The recorded number of counts is proportional to the 237U concentration 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

E = 1400000 O o 1200000 O 1000000 E 800000

600000 400000 200000

0

J

i l l

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.


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+) /d t - - 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 7 - O L ' 6 - t -

O 5 - E 4 - "E 3 - t~

2 - 1 -

0 " - 0

�9 _ I I

J

m i i

200 400 i i

600 800 1000

Time (hours)

Figure 4.13. Movement of the contaminant front through the column as measured by the 237U radiotracer (Experiment 1).


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

�9 E 0 4

103 t'- = 0 2 o 1 0

101

100

i - i - l l - - I . . . .

i i

0 200 400 600 800 1000 Time (hours)

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


2 5 0 0 0 0

2 0 0 0 0 0

t -

�9 ~ 1 5 0 0 0 0 , r

v

if)

o 100000 o

50000 t

39 1 - 1 8 h - 2 4 h - 4 2 h - 6 6 h - 9 0 h - 1 6 2 h ~ - 235 h - 3 3 1 h

9 - 4 2 9 h _ . .

A_ 1-} 1 7 " - . r 0 ' ~ . v . . I - - " - , ~ , " ~ , m , ~ ~ , ' , m I

- 5 0 5 10 15 20 25

He igh 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 Exper iment 1. The pseudo-f i rs t -order rate constant was es t imated 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

1 2 0 0 0 0 0 -

1 0 0 0 0 0 0 -

8 0 0 0 0 0 -

6 0 0 0 0 0 -

4 0 0 0 0 0 -

2 0 0 0 0 0 -

. . . . . R 2 = 0 .9957 _ _ _

J 0 I I

0 100 200 300 400

Time (hours)

500

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


16

1 4 -

1 2 -

~" 1o- o

,..4 8 - t - O :~ 6 - - - O n 4 - -

o

-2 o

1

y = 0 .0196x + 5.6681

R 2 = 0 .9756

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 237U signal, 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.


250000

200000 1- Oh ' " -" 9 2 - 1 3 h - - -

l 4- 93h 'E" + _ - / 7 / 7 - ~ 5 - 1 6 2 h _ __ �9 p__ 150000 - / / / / - ~ \ \ 6 -233 h ,,, / /L / / /~ 6 "~ \ \ 7 - 2 8 5 h ~= | / / ~ " \ \ \ \ 8 - 3 2 7 h

100000 . . . . .

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

" ' ' 6 0 100 200 3 0 4 0 500

Time (hours)

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.


16 .~ 14 E

12

.T= 10 o ' )

. i

~) 8 .,l:::

E 6 "5 4 0

2

. . S l l " - - - i ~

. _ I L

i i i i

O 100 200 300 400 500

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 half- lives, 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.


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 10%/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.


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 -50.03 (Johnson, 2000) have been obtained for log Ksp. In the review of Grenthe et al. (1992), log K s p - -48 .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 F.-G. Simon, V. Biermann

(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) 2- and 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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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 solution.

pH

pe

Activity of water

Ionic strength

Mass of water (kg)

Total alkalinity (eq/kg)

Total C02 (mol/kg)

Temperature (~

Electrical balance (eq)

Percent 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

1.217 X 10 -2

25.000 - - 1.842 X 10 -3

--3.52

ii

1.110264X102

5. 557905 X i0

0 D i s t r i b u t i o n o f species. ,.. Species ~. .

OH- H+

H2 0

C ( 4 ) 1 . 2 1 7 X HCO;

HzC03 MgHCOf C a H C O l NaHC03 CaCO,

MgC03

co4 -

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

NaCO; u02c03

C a ” + C a S 0 4 CaHCO: C a C 0 3 CaOH+

C a 4 . 5 0 1 X l o - ’

ci 3 . 3 6 4 x

uo2c1+ UCP+

c1-

Molal i ty A c t i v i t y L o g r n o l a l i t y L o g a c t i v i t y L o g gamma

1 . 2 1 4 x 1 . 1 5 8 X l o - ’ 5 . 5 5 1 X 1 0

9 . 8 7 1 X l o - ’ 1 . 8 4 9 X l o u 3

1 . 7 0 6 X l o A 4 4 . 5 5 9 x 9 . 1 7 9 X

6 . 3 4 5 X l o - & 1 . 9 5 3 x 1 . 5 9 5 x 8 . 4 2 5 X l o - ’ 3 . 0 2 8 X l o - ’

2 . 0 2 7 x 1 0 - ~

7 . 7 4 7 x l o +

3 . 2 2 6 x 1 0 - ~ 1 . 0 9 6 x 1 0 - ~

4 . 9 9 7 x l o - ’

1 . 7 0 6 X l o A 4 9 . 1 7 9 X

3 . 3 6 4 x 1 0 - ~ 3 . 4 4 0 x 3 . 1 4 9 X lo-*’

1 . 0 0 4 X lO-~’ 1 . 0 0 0 x l o - ’ 9 .993 x 1 0 - I

8 . 3 0 9 X l O 1 . 8 6 6 X l o p 3 1 . 6 8 5 X

1 . 4 4 3 X l o p 4 4 . 6 0 0 X l o - ‘ 9 .262 X l O - “

3 . 8 8 9 X

6 .402 X 9 . 4 9 1 X lo-’ 8 .892 X l o p 8

3 . 0 5 6 X LO-’ 7 . 0 9 2 x 10 -~

1 . 6 7 6 X

1 . 1 0 6 X l o - ’ 1 . 4 4 3 X l O 9 . 2 6 2 X l o p 6 4 . 2 2 7 X 1 0 -’

2 . 7 7 3 X 1 V 3 2 . 8 7 2 X

0 . 0 0 0

- 6 . 9 1 6 - 6 . 9 3 6

1 . 7 4 4

- 2 . 0 0 6 ~ 2 . 7 3 3 - 3 . 6 9 3 - 3 . 7 6 8 - 4 . 3 4 1 - 5 . 0 3 7 - 5 . 1 1 1 - 5 . 1 9 8 - 5 . 7 0 9 - 5 . 7 9 7 - 6 . 0 7 4 - 7 . 5 1 9

- 2 . 4 9 1 - 2 . 9 6 0 - 3 . 7 6 8 - 5 . 0 3 7 - 8 . 3 0 1

- 2 . 4 7 3 - 1 4 . 4 6 3 - 3 9 . 5 0 2

- 6 . 9 9 8 - 7 . 0 0 0 - 0 . 0 0 0

- 2 . 0 8 0 - 2 . 7 2 9 - 3 . 7 7 3 - 3 . 8 4 1 - 4 . 3 3 7 - 5 . 0 3 3 - 5 . 4 1 0 - 5 . 1 9 4 - 6 . 0 2 3 - 7 . 0 5 1 - 6 . 1 4 9 - 7 . 5 1 5

- 2 . 7 7 6 - 2 . 9 5 6 - 3 . 8 4 1 - 5 . 0 3 3 - 8 . 3 7 4

- 2 . 5 5 7 - 1 4 . 5 4 2 - 4 0 . 2 0 7

~ 0 . 0 8 2 - 0 . 0 6 4

0 . 0 0 0

- 0 . 0 7 5 0 . 0 0 4

- 0 . 0 8 0 - 0 . 0 7 3

0 . 0 0 4 0 . 0 0 4

- 0 . 2 9 9 0 . 0 0 4

- 0 . 3 1 3 - 1 . 2 5 4 - 0 . 0 7 5

0 . 0 0 4

- 0 . 2 8 4 0 . 0 0 4

- 0 . 0 7 3 0 . 0 0 4

- 0 . 0 7 3

- 0 . 0 8 4 - 0 . 0 7 8 - 0 . 7 0 5

z

3 s

a P -.

%’

2-

s P (b 3

%. a 9

(continued)

~~

Species Mo 1 a1 i t y A c t i v i t y L o g m o l a l i t y L o g a c t i v i t y

H ( 0 ) 1 . 4 0 3 X 1 0 - 2 5

H2

K+

K 1 . 2 8 2 X 1 0

KSO,

M g 4 . 4 5 2 X l O P 3 M g 2 +

MEIS04 MgHCO?;

MgC03 MgOH+

N a 1 . 1 9 9 X lop2 N a + N a S 0 ; N a H C 0 3 N a C 0 ;

O ( 0 ) 0 . 0 0 0

0 2

S ( 6 ) 9 . 0 7 7 X 1 0 so; ~

C a S 0 4

MgS04 N a S 0 ; KS04 HSO; uozso4

7 . 0 1 5 X lo-''

1 . 2 5 4 X l o r 4 2 . 7 9 3 X l O

3 . 2 5 9 x l o r 3 9 . 8 4 4 X 2 . 0 2 7 X l O ' 6 . 3 4 5 X lo-' 3 . 3 3 4 X lo-'

1 . 1 7 5 X l o - ' 1 . 8 9 7 X lo-' 4 . 5 5 9 X lop5 8 . 4 2 5 X

0.000

6 . 8 0 5 X 1 0 - 3

1 . 0 9 6 X 9 . 8 4 4 X

2 . 7 9 3 X 3 . 7 6 2 X l o p s 1 . 0 2 5 X

1 . 8 9 7 x 1 0 - ~

7 . 0 7 9 X

1 . 0 3 3 X 2 . 3 5 1 X 1 0 - 6

1 . 7 2 5 X lo-'

1 . 6 8 5 X 6 . 4 0 2 X l O 2 . 8 3 1 X l O *

9 . 9 3 3 x

9 . 8 3 8 X 1 . 5 9 7 X

7 . 0 9 2 X l o - ' 4 . 6 0 0 x 1 0 P

0.000

3 . 2 3 8 X

1 . 1 0 6 X l o p 3

1 . 5 9 7 X 2 . 3 5 1 X 3 . 1 4 2 X l O * 1 . 0 3 4 X

9 . 9 3 3 x l o p 4

- 2 5 . 1 5 4

- 3 . 9 0 2 - 5 . 5 5 4

- 2 . 4 8 7 - 3 . 0 0 7 - 3 . 6 9 3 - 5 . 1 9 8 - 7 . 4 7 7

- 1 . 9 3 0 - 3 . 7 2 2 - 4 . 3 4 1 - 6 . 0 7 4

- 4 2 . 0 8 5

- 2 . 1 6 7 - 2 . 9 6 0 - 3 . 0 0 7 - 3 . 7 2 2 - 5 . 5 5 4 - 7 . 4 2 5 - 1 1 . 9 8 9

Log gamma

- 2 5 . 1 5 0

- 3 . 9 8 6 - 5 . 6 2 9

- 2 . 7 6 3 - 3 . 0 0 3 - 3 . 7 7 3 ~ 5 . 1 9 4 - 7 . 5 4 8

- 2.007 - 3 . 7 9 7 - 4 . 3 3 7 - 6 . 1 4 9

- 4 2 . 0 8 1

- 2 . 4 9 0 ~ 2 . 9 5 6 - 3 . 0 0 3 - 3 . 7 9 7 - 5 . 6 2 9 - 7 . 5 0 3 - 1 1 . 9 8 5

0 . 0 0 4

- 0 . 0 8 4 - 0 . 0 7 5

- 0 . 2 7 6 0 . 0 0 4

0 . 0 0 4 - 0 . 0 7 1

- 0 . 0 8 0

- 0 . 0 7 7 - 0 . 0 7 5

0 . 0 0 4 - 0 . 0 7 5

0 . 0 0 4

- 0 . 3 2 3 0 . 0 0 4 0 . 0 0 4

- 0 . 0 7 5 - 0 . 0 7 5 ~ 0 .078

0 . 0 0 4

(continued)

Species

U ( 3 ) 0 . 0 0 0 u3 +

Molal i t y

2 . 0 5 3 X 10-I’ 5 . 9 9 2 X 1 0 - 3 5

1 . 9 8 0 X 1 0 ”

~

0 . 0 0 0

9 . 3 0 5 X

3 . 2 2 6 X

1 . 1 3 4 X

5 . 9 9 2 X l O 3 5

1 . 8 4 4 X 10- ’’ 3 . 1 4 9 X 1 0 4 3

0 . 0 0 0

1 . 4 2 7 x

1 .133 x

1 .980 x

4 . 5 5 7 x 1 0 -I4

1 . 9 5 3 X 1 0 1 . 5 9 5 X

3 . 0 2 8 X l O * 4 . 7 1 4 X lo-’’ 1 . 2 8 5 X

1 . 0 2 5 X 1 0 l2

~~~ ~

Activity Logmolality

9 . 9 7 5 X

6 . 0 4 7 X

9 . 6 2 2 x

0.000

7 . 7 6 9 X l O - l 8 3 . 2 5 6 X l o p z o I. 1 9 1 x

2 . 2 3 3 x

9 . 6 2 2 x 1 . 0 2 8 x

5 . 5 1 2 X

6 . 0 4 7 X

0.000 0.000

3 . 8 0 5 x 10-l~

9 . 4 9 1 X 1 0 ’ 8 . 8 9 2 X l I Y 8 3 . 0 5 6 X l o - * 3 . 9 3 5 X 1 0 - l ’ 6 . 2 4 2 X

1 . 0 3 4 X 1 0 - 1 2

- 1 2 . 6 8 8 - 3 4 . 2 2 2 ~ 3 5 . 7 0 3

- 5 1 . 0 7 9

- 1 7 . 0 3 1 - 1 9 . 4 9 1 - 2 2 . 8 4 6 - 2 6 . 9 4 5 - 3 1 . 9 4 6 - 3 4 . 2 2 2 - 3 5 . 7 0 3 - 3 7 . 7 3 4 ~ 3 9 . 5 0 2

- 1 3 9 . 8 1 4

- 1 3 . 3 4 1

- 5 . 7 0 9 - 5 . 7 9 7 - 7 . 5 1 9

- 1 0 . 3 2 7 - 1 1 . 8 9 1 - 1 1 . 9 8 9

Logac t iv i ty Log gamma

- 1 3 . 0 0 1 - 3 4 . 2 1 8 - 3 6 . 0 1 7

- 5 1 . 7 8 4

- 1 7 . 1 1 0 - 1 9 . 4 8 7 - 2 2 . 9 2 4 - 2 7 . 2 5 9 - 3 2 . 6 5 1 - 3 4 . 2 1 8 - 3 6 . 0 1 7 - 3 8 . 9 8 8 ~ 4 0 . 2 0 7

- 1 4 6 . 1 6 2

- 1 3 . 4 2 0

- 6 . 0 2 3 - 7 . 0 5 1 - 7 . 5 1 5

- 1 0 . 4 0 5 - 1 2 . 2 0 5 - 1 1 . 9 8 5

- 0 . 3 1 3 0 . 0 0 4

- 0 . 3 1 3

- 0 . 7 0 5

- 0 . 0 7 8 0 . 0 0 4

- 0 . 0 7 8 - 0 . 3 1 3 - 0 . 7 0 5

0 . 0 0 4 - 0 . 3 1 3 - 1 . 2 5 4 - 0 . 7 0 5 - 6 . 3 4 8

- 0 . 0 7 8

- 0 . 3 1 3 - 1 . 2 5 4

0 . 0 0 4 - 0 . 0 7 8 - 0 . 3 1 3

0 . 0 0 4

(continued) c.

Species M o l a1 i t y Act ivi ty Logmolality L o g a c t i v i t y L o g gamma

uo2 (SO4 12- 2 . 0 5 3 x 9.975 x 1 0 - l ~ - 12.688 - 13.001 - 0 . 3 1 3 uo2c1+ 3 . 4 4 0 X 10- l5 2.872 X 10 - 14.463 - 14.542 - 0.078 (UO,)Z(OH)~+ 1.921 X 10 -lE 9.332 X - 15.717 - 16,030 - 0.313 ( UOz ) 3 (OH ) 8.179 X 6 . 8 2 8 X lo-'* - 17.087 - 17.166 - 0.078


Saturation indices.

Phase SI log IAP log KT

Anhydrite - 0.63 -- 5.27 - 4.64 CaS04

Aragonite 0.15 -- 8.19 -- 8.34 CaC03

Artinite -- 6.54 3.06 9.60 MgC03 "Mg(OH)2 : 3H20

B_U02 (OH) 2 -- 3.75 1.79 5.54 UO2(OH)2

Brucite -- 5.56 11.24 16.79 Mg(OH)2

Calcite 0.29 -- 8.19 -- 8.47 CaC03 CH4(g) -- 67.31 -- 107.41 -- 40.10 CH4

C02 (g) -- 1.25 -- 19.41 -- 18.16 C02

Dolomite 0.64 -- 16.36 -- 17.00 CaMg(C03)2

Epsomite -- 3.12 -- 5.26 -- 2.14 MgSO4:7H20

Gummite -- 8.61 1.79 10.40 U03

Gypsum -- 0.42 -- 5.27 - 4.85 CaS04 : 2H20

H2(g) -- 22.05 -- 22.00 0.04 H2

Halite - 6.15 - 4.56 1.58 NaCl

Huntite -- 2.74 -- 32.71 -- 29.97 CaMg3(C03)4

Hydromagnesite -- 12.69 -- 21.46 -- 8.77 Mgs(CO3)4(OH)2 : 4H20

Lime -- 21.57 11.22 32.80 CaO

Magnesite -- 0.14 -- 8.17 -- 8.03 MgC03

Mirabilite -- 5.39 - 6.51 - 1.11 Na2SO4-lOH20

Natron -- 8.12 --9.43 -- 1.31 Na2CO3:lOH20

Nesquehonite - 2.55 - 8.17 - 5.62 MgC03 �9 3H20

02(g) -- 39.12 44.00 83.12 02

Periclase -- 10.27 11.24 21.51 MgO

Portlandite - 11.45 11.22 22.68 Ca(OH)2

Rutherfordine - 3.15 - 17.61 - 14.46 U02C03

Schoepite - 3.61 1.79 5.40 UO2(OH)2 :H20

Thenardite - 6.32 - 6.50 - 0.18 Na2S04

Thermonatrite -- 9.55 -- 9.42 0.13 Na2C03 :H20

U308 (C) -- 10.07 -- 16.61 -- 6.54 U308

U409 (C) -- 18.57 -- 58.82 -- 40.25 U409

U02 (am) -- 11.92 -- 20.20 -- 8.28 U02

U03(C) -- 5.92 1.79 7.72 UO3

Uraninite -- 6.29 -- 20.20 -- 13.92 U02


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,

Laboratory tests using natural groundwater 113

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 Ca Mg H C O 3 804 CI Eh conductivity (Ixg/l) (mg/l) (mg/l) (mg/l) (mg/l) (mg/l) (mg/l) (mV) (IxS/cm)

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 .23 11,253 18.3 12,604 380 1749 254

361 43 250 43 128

85 170 85 115

1455 167 1313 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 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)

Influent a 7.11 1496 718 1030 168 53 474 361 43 250 250 Effluent 8.20 1272 17 867 82 60 367 421 50 2.6 234

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


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

03 "-1

v

d c -

O L )

E t "

t _

1000

100

10

- rn rnt~

.= v

0 100

A.= A V v v

2+0 3+0 460 5+0 6+0 700 Bed volumes

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 uranium- contaminated groundwater: selected groundwater and column effluent constituents (average values, duration of experiments 6 months).

Parameter pH Electrical conductivity U Ca Mg HCO3 (IxS/cm) (Ixg/l) (mg/l) (mg/l)

Influent a 7.46 1726 913 190 64 630 Effluent column 1 7.99 1539 < 10-900 b 142 62 522 Effluent column 2 8.06 1463 < 10-160 c 108 63 414

a 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).

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

Laboratory tests using natural groundwater 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.

Experimental Test site Unit

Na 275 44 mg/l K 5 14 mg/l Ca 180 140 mg/l Mg 108 40 mg/l CI 119 32 mg/l $04 870 253 mg/l C03 10 10 mg/l HCO3 741 412 mg/l Carbonate hardness 34 18.9 -~ a Total hardness 50.4 28.8 -~ pH 7.2 6.8 U 850 385 mg/l Electrical conductivity 2433 1061 IxS/cm

a 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

/

o o ;,. "o o o~176 ~o'~ P/ ,~Yo ~ I

, ~ ,:. ~ ~ o % . . o o . ,o OoOo~ ,%o0,

~ .~2~~ _ _ J -o %, o., ,uo ',: 0 ,.,.-o,~ , ~'~o ~ 9 o ' , , - o oHiO... ~ .~ ?, ),%~_ / o ~%0,,02 o o ,%~176 ~o_,:, o'<-d.~ o o, ,Y,=~azic

~-A o ~ ~ ,%.0 O" ,,

/~u t,v koA1 /

Figure 5.4. Schematic drawing of a test box.

118 J. B r h m et al.

Figure 5.5. Experimental box: full view (left) and inside view (fight).

a more stable long-term hydraulic conductivity. The hydroxyapatite was a very fine white powder, so to increase its hydraulic conductivity 20% hydroxyapatite was mixed with 80% quartz sand by volume. In the experiment with steel fibres sand was used to fill the pores. The dimensions of the filled boxes are compiled in Table 5.5, and Figure 5.6 shows photographs of the filled boxes.

The operation of the test boxes required a volume of about 0.8 m 3 of water per day for each box to be fed to the system. As the flow through the test boxes was to be achieved by gravity, three I m 3 buffer tanks were placed 7 m above the test boxes. The buffer tanks were refilled daily with the effluent from the mine water treatment station. This procedure offered a highly independent and simple feeding system. Figure 5.7 illustrates the set-up for the box tests on the premises of Mecsek Ore Environment Co. in Prcs. Driven by a hydraulic gradient, water was directed from the buffers into the overflow system of the boxes. After passing through the box and the reactive material filling, water was directed into a collector while passing several measurement devices and sampling ports. Each test box had four sampling ports. The maximum flow velocity was 0.55 l/min at the beginning of the tests.

Coefficients of permeability kf were determined on additional samples as follows: a closed cylinder with known geometric parameters was filled with the material. Then the overflow of water from a pre-set height was measured within a time interval, and kf was calculated based on Darcy's law (kf = Q/ (A i) with Q - flow rate of overflow water (m3/s), A = area of the cylinder cross-section perpendicular to the flow direction (m2), and

i = hydraulic gradient (m/m)). The results of these preliminary tests are presented in Table 5.6.

Table 5.5. Dimensions of the test boxes used for the experiments.

Overall Thickness of Width Height Bed volume of thickness (m) active zone (m) (m) (m) active zone a (m 3)

Iron-matrix 0.8 0.2 1 0.8 0.16 Steel-matrix 0.8 0.6 b 1 0.8 0.48 HAP-matrix 0.8 0.2 l 0.8 0.16

a Volume of the reactive matrix. The pore volume can be estimated by multiplying it by the porosity. b The steel fibres were difficult to handle and it was impossible to place them with a width of only 0.2 m.


Figure 5.6. Sand/iron mixture (left), steel fibres (center) and HAP (right) in the test boxes.

The most important experimental data such as hydraulic gradient, pH, Eh, uranium, iron, oxygen, calcium, magnesium and bicarbonate concentrations, temperature and the flow rate of the water overflow were measured in the feed and the outflow during the experiment. After completion of the experiments, the reactive media were examined using micro-probe and scanning electron microscopy (SEM) analysis.

2. Experiments on shredded cast iron

A total of 264 bed volumes (m 3 of water/m 3 volume of reactive matrix) of uranium- contaminated water was passed through the experimental box filled with the iron/sand mix (240 kg iron and 192 kg sand). The initial flow rate through the box was 0.55 l/min at a hydraulic gradient of 0.015 m/m, reducing to 0.012 l/min at 0.03125 m/m at the end of the experiment. Thus, the permeability of the experimental system changed from kf -- 7.64 x 10 -4 to 8.07 • 10 -6 re]s, a value two orders of magnitude lower than at the beginning of the experiment. At the end of the experiment, reactive material was sampled from the box, and non-oxidised iron particles were detected only 20-22 mm behind the

Figure 5.7. The experimental set-up of the floor-scale test boxes and its schematic sketch.

120 J. Brhm etal.

Table 5.6. Coefficients of permeability kf of the reactive media and the sand used in the box experiments.

Material Coefficient of permeability kf (ITI]S)

Iron particles 1.58 • 10 - 4

Sand 8.57 x 10 -4 Iron/sand mix 3.90 x 10 -5

(50%/50% by volume)

up-gradient interface to the sand filtration layer. Iron corrosion has obviously occurred predominantly in a narrow layer at the up-gradient interface. This observation was verified by the SEM analysis.

Water chemistry and other parameters verified the results of previous (column) investigations. The pH of the percolated water increased during the experiment due to Fe ~ oxidation and water reduction (see Chapter 3). Moreover, the Eh was always below 200 mV favouring uranium (VI) to uranium (IV) reduction

Fe + uoZ+(aq)=~Fe 2+ + UO2(s ) (5.~)

Reduction of UO 2+ takes place at the conditions prevalent in the box experiments. Solubility of U02 is low between pH 4 and 14 and the reduced species precipitates. Therefore, it is important that the Eh value is kept below the U(IV)/U(VI) limit curve (see Chapter 4).

Although uranium reduction and precipitation is the important mechanism for removing uranium from solution, uranium compounds are only a very small proportion of the precipitates formed in an elemental iron barrier. The primary cause of precipitates within the reactive zone is the increase in pH caused by Fe~ oxidation, which together with the dissolved oxygen and carbonate in the water encourages precipitation of minerals such as Fe(OH)e, FeCO3 and CaCO3. In the box test, formation of these precipitates gradually filled the pores of the iron/sand mixture, reducing its hydraulic conductivity.

SEM images and micro-probe results verified the initial assumption that a corrosion front moves through the reactive material. At the end of the box experiment, this front had reached 20-25 mm across the 2OO-mm wide iron/sand layer over the entire cross-section of the box. Uranium was found within this section only. The investigation showed that uranium was mainly co-precipitated with other minerals: uranium was mainly detected within the pores (Figs 5.8 and 5.9).

Examination of different cross-sections of the reactive iron/sand mix layer perpendicular to the water flow provided detailed information on the processes inside the reactive layer during the experiment. The entry point of the contaminated water into the reactive layer (up-gradient interface) was chosen as the first cross-section. The pores were here clearly filled with Fe(OH)e, FeCO3, CaCO3 and iron oxide (Fig. 5.10). These precipitates completely clogged the pores and decreased the hydraulic conductivity of the reactive layer significantly. Iron particles were oxidised and most probably lost their reactivity. Similar composition was found at cross-sections 15 and 20 mm from the


Figure 5.8. Backscatter and morphology images of uranium precipitation in the iron matrix.

up-gradient interface (Figs 5.11 and 5.12). At 25 mm from the up-gradient interface clogging of the pores is not as intensive as at the locations before (Fig. 5.13). SEM images of cross-sections 30 mm from the up-gradient interface show only minor clogging of pores, with precipitates of Fe(OH)2, FeCO3 and CaCO3 being present mainly as thin films on the iron and sand particles (Fig. 5.14).

3. Geochemical modelling of mining water in contact with elemental iron

Results from geochemical modelling (using the PHREEQC code with the Minteqa2 database) of the water used for the experiments in the box tests have been presented in

Figure 5.9. Spectrum of uranium precipitation in the iron matrix.

122 J. B r h m et al.

Figure 5.10. First cross-section through PRB (starting position). (1) sand particle, (2) mainly precipitated CaCO3, (3) iron particle. The surface is covered with Fe(OH)2 and FeCO3 precipitation.

Chapter 4. To perform geochemical simulation runs for water in contact with elemental iron, all the species that may be involved have to be in the database of the program. Therefore, elemental iron was inserted in the database as a new phase (Fig. 5.15). The required data were retrieved from the database of the computer code Geochemist ' s Workbench (Johnson, 2000).

The input file for the simulation run is displayed in the Appendix 5A. Iron is entered as an equilibrium phase (Fe O 0.05, the first number is the target saturation index, the second is the amount in moles). Oversaturated compounds have been allowed to precipitate. This is achieved by insertion of a new block "solid solution" in the input file and listing of the respective minerals. The results of the simulation run are listed in Appendix 5B.

Figure 5.11. SEM images of the cross-section at 15 mm distance from the starting position. Precipitates and uranium are present.


Figure 5.12. SEM images of the cross-section at 20 mm distance for the starting position. Similar composition of precipitates and uranium are present.

The results from the simulation run show that under the conditions prevailing in the box

experiments with elemental iron uranium is reduced to the oxidation state IV and

precipitated as uraninite. Sulphate is almost completely reduced to sulphide. Formation of

hydrogen is indicated by the positive saturation index in Table A5.lO (Saturation Indices) at the end of the results. The pH is in the alkaline region. Under equilibrium conditions the precipitated mass (excluding iron phases) amounts to more than 800 mg/l solution treated as displayed in Table 5.7.

Figure 5.13. SEM image of the cross-section at 25 mm distance from the starting position. Clogging of pores is not as intensive as in previous cross-sections. Dominant precipitations are Fe(OH)2, FeCO3 and CaCO3. Uranium is detectable, but not dominant. This position is possibly the location of the front.

124 J. B#hm et al.

Figure 5.14. SEM image of the cross-section at 30 mm from the starting position. Only minor clogging of pores can be observed. Precipitation of Fe(OH)2, FeCO3 and CaCO3 as thin films on the particles, no uranium was detected. Iron particles are less corroded.

4. Experiments on waste steel fibres

During the experiment using waste steel fibres (80 kg in total), 135 bed volumes of uranium-containing water were treated. The flow rate measured at the exit overflow was 0.51 l/min at a hydraulic gradient of 0.0025 m/m at the start of the experiment, decreasing to 0.25 l/min at 0.0125 m/m by the end of the experiment. Thus, the permeability of the experimental system decreased from kf-- 1.70 x 10 -3 to 4.66 x lO -4 m/s, which is a decrease of nearly one order of magnitude. Water chemistry data showed that similar processes occurred with the steel fibres as in the iron/sand matrix. Therefore, it can be concluded that uranium precipitation and other precipitation processes were similar in the

experiments on waste steel fibres to those using granular iron particles as a reactive

material. However, the permeability decrease was smaller.

Fe Fe +2.0000 H+ +0.5000 02 = + 1.0000 Fe+2 + 1.0000 H2O Iog_k 59.0325 delta_H -88.87 kcal/mol # Calculated enthalpy of reaction in kJ

# -delta_H -372.029 kJ/mol # Calculated enthalpy of reaction Fe # Enthalpy of formation: 0 kcal/mol # -analytic-6.2882e+001 -2.0379e-002 2.0690e+004 2.3673e+001 3.2287e+002 # -Range: 0-300 # inserted from Ilnl.dat

Figure 5.15. Database entry for elemental iron as an additional phase.


Table 5.7. Composition of precipitated fraction as a result of geochemical modelling using PHREEQC.

Mineral Amount (mol) Molecular weight Mass (g) (g/mol)

Calcite 1.39 x 10 -4 216.88 0.0301 Aragonite 1.01 x 10 -4 216.88 0.0219 Uraninite 3.58 x 10 -6 270.03 0.0010 Dolomite 4.26 x 10 -3 183.92 0.7855 Sum 0.8386

Unfortunately, it was not possible to perform SEM and micro-probe analyses, due to the structure of the fibres. Uranium concentrations were below 50 Ixg/l in the overflow water during operation, pH and Eh data were in the range of possible U(VI) reduction to U(IV). Uranium breakthrough was not detected in the box effluent.

5. Experiments on hydroxyapatite

In the box experiment with HAP as a reactive material (11.5 kg HAP and 307 kg sand), 340 bed volumes of water containing uranium were treated. The flow rate measured at the exit overflow was initially O. l l/min at a hydraulic gradient of 0.02 m/m, but this increased rapidly at the start of the test to 0.45 l/min at 0.01875 m/m, and then remained fairly constant until the end of the test. Thus, there was a rapid increase in permeability of the experimental system from kf = 4.20 x 10 -5 to 5.0 x 10 -4 rn]s (i.e. about one order of magnitude).

The uranium concentration in the effluent increased steadily after treatment of 50 bed volumes reaching a full breakthrough (Fig. 5.16) after 150-200 bed volumes. The early

800

700

600 :::L v 500 6 t - O o 400 E

�9 - 300 t '- t~

D 200

100

A A zx

A A

�9 HAP effluent

A HAP feed

~ a . a ~ o

w

A �9 ,~ �9 oo ~ 1 7 6 _ A - ~ , , A o ~

I l i

0 5b 1;0 150 2;o 2;o a00 35o Bed volumes

Figure 5.16. Uranium concentration in the effluent of the test box using HAP as a reactive material.

126 J. B6hm etal.

1.2

o

�9 9- ~ 0.6 0

1.0 .,....../_ m i , - - " '~- ' - ' "~ ~ _l ~mm �9

0.8 / �9 �9 �9

�9 �9

�9 �9

0.4

0.2 / �9 c/c 0column effluent m

0.0 ~ � 9 U sorbed 0

0 1000 2000 3000 4000 5000

I/kg HAP

- 600

500

03 400 " 03 E v

300 -o _(3

0

200

100

Figure 5.17. Ratio of outlet (c) and inlet (Co) concentrations as a function of the treated water volume per kg of HAP. The cumulative uranium sorbed per kg of HAP (calculated from C/Co) is depicted as solid line.

breakthrough of the uranium can be explained by the low concentration of the reactive material HAP in the box experiments, which was by far lower than in the experiment using

elemental iron (11.5 kg HAP = 22.9 mol HAP = 68.7 mol phosphate compared to 240 kg i r o n - - 4 3 0 1 mol Fe). Plotting the volume of treated water per kg of HAP against the

concentration ratio C/Co (Fig. 5.17) indicates that the 50% breakthrough (i.e. C/Co = 0.5) occurs between 600 and 1200 1/kg HAP. The maximum sorption capacity of the reactive

material is around 5 0 0 m g U / k g HAP (also shown in Fig. 5.17). This value is

Figure 5.18. SEM image of the HAP/sand matrix: no other material but sand (1) and hydroxyapatite (2) were detected. The pores are clean and not clogged by any precipitates.


approximately a factor of 6 lower than the results from the column experiment using activated uranium as a tracer (see Chapter 4). Fuller et al. (2002) recorded breakthrough curves for different phosphate rocks and apatite materials. After sorption of some hundreds of mg/kg of uranium onto phosphate rock uranium breakthrough occurred indicating that the sorption capacity of the reactive material was exhausted. With fertiliser-grade bone meals several thousand mg/kg sorbed uranium have been reached.

SEM images of the HAP/sand mix failed to detect any other precipitation products after the end of the experiments (Fig. 5.18). However, the detection limit of the analytical system might have been too high to detect modest precipitate concentrations. Fuller et al. (2002) found in their work that uranium was not detected either in backscatter image or by EDS in HAP with 4700 ppm U(VI). At higher uptake levels (77,500 ppm) they found a strong backscatter emission indicative of a uranium-bearing phase. As no uranium was detected in the present study by SEM analysis, it is difficult to reach a definite conclusion about the uranium attenuation mechanism of HAP in the box experiment.

6. Long-term performance of the reactive materials

Changes in reactivity and hydraulic properties are the most important parameters for predicting the long-term performance of different reactive materials. Figure 5.19 shows the removal of uranium in all three test box experiments. Iron particles and steel fibres showed very good removal efficiencies and result in particularly low uranium concentrations in the effluent (generally less than 50 txg/l), but uranium removal by HAP decreased after about 50 bed volumes indicating that the maximum retention capacity of the reactive material had been reached. This seemed to be the main problem, though the HAP concentration in the sand was only 3.6% by mass. The increase in pH in the steel and the iron systems resulted in a significant reduction in the calcium concentration (Fig. 5.20). This finding coincides with the results of geochemical modelling (see Section 3).

1000 900 A HAP [] Steel o Iron l--

�9 Feed_Steel �9 Feed_HAP �9 Feed_Iron / 800 .,, .

700 �9 " in �9 �9

c5 600 �9 �9 �9 �9 �9 h : " ~ , _ - ' b " _ . , ' . ,_ . .

o 5 0 O r l ~ _ . . 7 . . , . ~ , , ," _ ~ . x t . ~ - . . �9 E= 400 ---A- �9 " " �9 �9 - . �9 A

" " " : ' ' n m ~ " ~ ~ " " A #e~ . �9 300 u - ,m, "- A " ~ �9 A

Oo ~ A A - � 9 1 4 9 �9 �9 2 0 0 ~ A ~ ~�9 �9 ~ �9

1 0 0 ~ ~ ; ; i n A a ~D~. o

........... o o i

0 50 100 150 200 250 300 350

Bed volumes

Figure 5.19. Attenuation of uranium concentration using different matrix materials.

128 J. Brhm etal.

&

200 . . . . i l l i [ m m l m m m m ~ m m [ m m

160 l i " i I ;,A~-Z/' -'-! . . . . ~x-~ . . . . . . -~ �82 ~ - -~A . . . . . . . . - ~ - -

~ 140 - - i . . . . . . . . . . . . . . . . . . . . . . . . . . . .

o= 120 ~ - - - m_ _ ~ _m.d~ "- ~ Zx HAP [] Steel o iron ~ tJ D D D ~ I ~ .Feed Steel �9 HAP .Feed Iron ~ o 1OO D�9 - - -

IF [] [] = 80 [] [] OOo o

t~ 60 ~ - - ~ 1 7 6 ~oe3o O .... _(q .oO o- 40 o 40

20

O I I i I I

0 50 100 150 200 250 300 Bed volumes

Figure 5.20. Calcium concentration in the experiments in the test boxes.

350

The permeability of all three reactive materials in the test boxes showed distinct trends over the experiment (Fig. 5.2 l). After a rapid increase at the beginning of the experiment, the permeability of HAP was constant during the experiment. The reasons for this behaviour are not fully understood. There is a chance that the finest particles were washed out of the HAP/sand matrix at the beginning of the experiment. However, more likely is the assumption that preferential flow pathways were formed. This may also explain the lower sorption capacity of only 500 mg U/kg HAP, as compared to the column experiments.

Loss of permeability was observed in both iron-based systems. By assuming exponential decay behaviour (a straight line in a semi-logarithmic graph), it is possible to predict how many bed volumes can be treated until the kf value of the reactive material reaches that of the surrounding soil, which in a full-scale application may lead to a

1 E 0 2

I Zx HAP [] Steel o Iron I

E l E O 3

- - O t ~

E $ l E O 4

&

l E O 5 O 50 100 150 200 250 300 350

Pore volumes 400

Figure 5.21. Permeability of the matrixes as a function of bed volumes treated.


Figure 5.22. Picture of iron sponge used in commercial PRB installations.

hydraulic by-passing of the barrier. Precipitation occurred at the cross-section perpendicular to the flow direction, where water met non-oxidised iron particles in the iron-based reactive materials (precipitation front). Most of the newly formed minerals caused clogging of the pores. Uranium precipitates mainly in this section of the reactive layer.

In both iron-based systems the permeability is decreasing with time. The precipitation front moves parallel with the flow. In this case, the clogging of pores is uniform across the whole cross-section. Clogging impairs the function of the barrier although non-reacted iron remains there. Building wider barriers is, therefore, not a solution enhancing the lifetime of a barrier. New materials or material mixtures have to be developed to overcome this problem. In a commercial project with PRBs in Germany, iron sponge is used instead of shredded cast iron (Ebert et al., 1999). This material is pelletised and usually applied without adding sand as an inert material. The pores are much larger and a failure of the system due to clogging of pores has not yet been observed. Figure 5.22 shows a picture of iron sponge.

D. Conclusions

Performance of reactive barriers is strongly limited by the hydraulic properties of the reactive media and chemical characteristics of the accompanying water. On the example of two different iron materials - shredded grey cast iron and waste steel fibres - it has

130 J. Br"hm et al.

been shown in laboratory experiments with uranium-contaminated groundwater that an elemental iron barrier causes precipitation of iron and alkaline earth carbonates, which leads to clogging of pores and subsequently to a reduced permeability. Although the principal uranium attenuation mechanisms appear to be very efficient leading to an almost complete removal of the dissolved uranium from the contaminated water, these precipitation reactions may impair the function of a reactive barrier system. High Ca and Mg concentrations as present in the groundwater of Prcs will have an influence on the long-term performance of iron-based PRB systems. The application of hydroxyapatite (HAP) does not cause major changes in water chemistry. However, the removal capacity seems to be limited. Breakthrough was observed in several experiments using HAP as a reactive material.

Interactions between groundwater constituents and the reactive media have to be further examined to better understand the hydrochemical processes taking place in PRB systems. Laboratory investigations on elemental iron materials showed that the bulk of the precipitates form at or near the upstream interface between iron material and soil (or filter sand/gravel). The extent of such precipitation reactions is strongly dependent on basic hydro/geochemical conditions such as pH, Eh and composition of the groundwater. Although great care was taken in the experiments to reproduce natural conditions as closely as possible, especially by using natural groundwater taken from the former uranium mining and processing site near Prcs, Hungary, some parameters, especially Eh and temperature, could not be fully controlled. Experiments on the site or even within the aquifer itself would therefore be advantageous to overcome the limitations of laboratory experiments where the feed water needs to be transported from the site to a laboratory. Approaches to perform such experiments are described in Chapter 6.

References

Ebert, M., Mrller, W., Wegner, M., 1999. R and D Project: permeable reactive barrier (PRB) in Rheine- latest results. Altlastenspektrum 2/99, 105-112.

Fuller, C.C., Piana, M.J., Bargar, J.R., Davis, J.A., Kohler, M., 2002. Evaluation of apatite materials for use in permeable reactive barriers for the remediation of uranium-contaminated groundwater. 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.

Johnson, J., 2000. Geochemist' s Workbench. Lawrence Livermore National Laboratory, Database thermo. com, V8R6230.


Appendix 5A. Input file for PHREEQC geochemical simulation of uranium-contaminated water in contact with elemental iron

TITLE reactive barriers and uranium

SOLUTION 1

temp 25 pH 7

pe 4

redox pe

units mg/l

density 1

U 0.85 Ca 180

S(6) 870 as SO4 C(4) 741 as HC03

Cl 119

Na 275 K 5

Mg 108

-water l # kg SAVE solution l END USE solution l

EQUILIBRIUM_PHASES 2

Fe 0 0.05

SOLID_SOLUTIONS 1

Precipitate

-comp -comp

-comp -comp -comp -comp

Calcite 0 Aragonite 0 Uraninite 0 Dolomite 0 Pyrite 0 Hematite 0

END

132 J. Bb'hm et al.

Appendix 5B. Results of a simulation run of a PHREEQC geochemical simulation of uranium-contaminated water in contact with elemental iron.

Description of solution.

pH

pe

Activity of water

Ionic strength

Mass of water (kg)

Total alkalinity (eq/kg)

Total CO~ (mol/kg)

Temperature (~

Electrical balance (eq)

Percent error, I00 X (Cat -- IAnl) /(Cat + IAnl)

Iterations

Total H

Total 0

i0. 933 Charge balance

- ii. 840

0.999

1.735 X I0 -~

9. 993 X 10 -I

1.099 X i0 -~

3. 412 X 10 -3

25.000

- 1.842 X 10 -3

-7.14

5O

1.110264 X 102

5. 547787 X 101

Distribution of species.

Species Molality Activity Log Log Log

molality activity gamma

OH- 9.836 X 10 -4 8.601X 10 -4 H + 1.304 • 10 -11 1.167 • 10 -11

H20 5.551• i0 +l 9.989 • 10 -I

C(4) 3.412 X 10 -3

CO~- 2.552 • 10 -3 1.547 • 10 -3

HCO~ 4.369 • 10 -4 3.855 X 10 -4

NaC03 3.332 X 10 -4 2.940 • 10 -4

MgC03 8.687 X i0 -S 8.722 • I0 -S

NaHC03 2.216 X 10 -6 2.224 • 10 -6

MgHCO~ 3.055 X i0 -v 2.678 • i0 -v

CaC03 2.233 • 10 -8 2.242 X 10 -8

H2C03 1.006 • 10 -8 1.010 • 10 -8

CaHCO~ 4.606 X 10 -11 4.075 X 10 -11

U02 ( C03 ) 4- 3 2.219 • i0 2s 2.755 • 10 -29

U02 (C03)~- 1.246• -3o 7.396Xi0 -31

U02C03 5.964 • i0 -3s 5.988 • -35

-- 3.007 -- 3.065 -- 0.058

- - 10.885 -- 10.933 -- 0.048

1.744 -- 0.000 0.000

-- 2.593 -- 2.811 -- 0.218

- - 3.360 -- 3.414 -- 0.054

-- 3.477 -- 3.532 -- 0.054

- - 4.061 -- 4.059 0.002

-- 5.655 -- 5.653 0.002

-- 6.515 -- 6.572 -- 0.057

-- 7.651 -- 7.649 0.002

-- 7.997 -- 7.996 0.002

-- 10.337 -- 10.390 -- 0.053

- - 27.654 -- 28.560 -- 0.906

-- 29.904 -- 30.131 -- 0.227

-- 34.224 -- 34.223 0.002

Ca 3. 918 X 10 -8

CaC03 2.233 X 10 -8 2.242 • -8 -- 7.651 - 7.649 0.002

Ca 2+ 1.655 X 10 -8 1.020 X 10 -8 -- 7.781 -- 7.991 - 0.210

CaOH + 2. 492 X 10 -I~ 2. 204 • 10 -I~ -- 9. 603 -- 9. 657 -- 0.053

CaHCO + 4.606 • 10 -11 4.075 X 10 -11 -- i0.337 -- i0.390 -- 0.053

CaS04 0.000 0.000 -- 40.300 -- 40.298 0.002


(continued)

Species Molality Activity Log Log Log

molality activity gamma

Cl 3.367 X 10 -3

Cl- 3.367 X10 -3 2.938 X 10 -3 - 2.473 -- 2.532 -- 0.059

FeCl 2+ 2.136 X 10 -36 1.285 X 10 -36 -- 35.670 -- 35.891 -- 0.221

FeCl~ 1.915 X 10 -38 1.687 X 10 -38 -- 37.718 -- 37.773 -- 0.055

FeC13 0.000 0.000 -- 41.307 -- 41.305 0.002

U02Cl + 0.000 0.000 -- 43.767 -- 43.824 -- 0.057

UCl 3+ 0.000 0.000 -- 53.032 -- 53.542 -- 0.510

Fe(2) 1.334 X 10 -8

Fe (OH) 3 7. 682 X 10 -9 6. 765 X 10 -9 -- 8. 115 -- 8. 170 -- 0. 055

FeOH + 3.314 X 10 -9 2.919 X 10 -9 -- 8.480 -- 8.535 -- 0.055

Fe (OH) 2 2. 118 X 10 -9 2.127 X 10 -9 -- 8. 674 -- 8. 672 0. 002

Fe 2+ 1. 761 X 10 -I~ 1. 078 X 10 -I~ -- 9. 754 -- 9. 967 -- 0. 213

Fe (HS)2 5. 163 X i0 -11 5. 183 X 10 -11 -- i0. 287 -- I0. 285 0. 002

Fe (HS)3 1. 494 X 10 -13 1. 311 • 10 -13 -- 12. 826 -- 12. 882 -- 0. 057

FeS04 0.000 0.000 -- 42.335 -- 42.333 0.002

Fe(3) 2.218 X 10 -13

Fe (OH)[ 2. 216 X 10 -13 i. 955 X 10 -13 -- 12. 654 -- 12. 709 -- 0. 054

Fe(OH)3 2.275 X I0 -z6 2.284XI0 -z6 -- 15.643 -- 15.641 0.002

Fe (OH) + 2. 573 X 10 -19 2. 270 X 10 -19 -- 18. 590 -- 18. 644 -- 0. 054

FeOH 2+ 1. 331 X 10 -26 8. 008 X 10 -27 -- 25. 876 -- 26. 096 -- 0. 221

Fe 3+ 3.941X 10 -35 1.449 X 10 -35 -- 34.404 -- 34.839 -- 0.435

FeCl 2+ 2.136 X i0 -36 1.285 X i0 -36 -- 35.670 -- 35.891 -- 0.221

FeCI2 + 1. 915 • 10 -38 1. 687 X 10 -38 -- 37. 718 -- 37. 773 -- 0. 055

FeCI3 0. 000 0. 000 -- 41. 307 -- 41. 305 0. 002

Fe2 (OH)24+ 0. 000 0. 000 -- 49. 857 -- 50. 763 -- 0. 906

FeSO + 0.000 0.000 -- 65.480 -- 65.535 -- 0.055

Fe3 (OH) 5+ 0. 000 0. 000 -- 65. 671 -- 67. 087 -- 1. 416

Fe(SO4)2 0.000 0.000 -- 98.594 -- 98.650 -- 0.057

H(O) 9.172 X 10 -2

H2 4. 586 X 10 -2 4. 604 X 10 -2

K i. 283 X 10 -4

K + i. 283 X 10 -4 i. 119 X 10 -4

KS04 2. 160 X 10 -38 i. 906 X 10 -38

Mg 1.915 X 10 -4

Mg 2+ 9.496 X 10 -5 5.910 X 10 -5

MgCO3 8.687 X 10 -5 8.722 X 10 -5

MgOH + 9.375 X 10 -6 8.310 X 10 -6

MgHCO~ 3.055 X 10 -7 2.678 X 10 -7

MgSO4 2.536 X 10 -37 2.546 X 10 -37

Na 1.200 X 10 -2

Na + 1.166 • 10 -2 1.025 • 10 -2 _

NaC03 3.332 X 10 -4 2.940 X10 -4

NaHC03 2.216 X10 -6 2.224 • -6

NaSO[ 1.411X10 -36 1.245 X 10 -36

o(0)o.o0o 02 O. 000 O. 000

-- 1. 339 -- 1. 337 0.002

- - 3.892 -- 3.951 -- 0.059

-- 37.666 -- 37.720 -- 0.054

- - 4.022 -- 4.228 -- 0.206

- - 4.061 -- 4.059 0.002

-- 5.028 -- 5.080 -- 0.052

-- 6.515 -- 6.572 -- 0.057

- - 36.596 -- 36.594 0.002

- - 1.933 -- 1.989 -- 0.056

-- 3.477 -- 3.532 -- 0.054

-- 5.655 -- 5.653 0.002

- - 35.850 -- 35.905 -- 0.054

-- 89.709 -- 89.707 0 . 0 0 2

134 J. B 6 h m et al.

(continued)

Species Molality Activity Log

molality

Log

activity

Log

gamma

S(--2) 8.924 X 10 -3

s~ s[- S~- HS-

S 2-

s~ S~- H2S

Fe (HS) 2 _

Fe (HS) 3

8.522 X 10 -4

4.972 X 10 -4

4.411 X 10 -4

2.656 X 10 -5

3.997 X i0 -v

1.752 X i0 -v

9.944 • 10 -9

2.360 X 10 -9

5.163 X i0 -11

1.494 X I0 -13

5. 059 X 10 -4

2. 952 X 10 -4

2. 619 X 10 -4

2. 323 X 10 -5

2. 405 X 10 -v

1. 040 X 10 -v 5. 903 • 10 -9

2.369 X 10 -9

5. 183 X i0 -11

1.311X i0 -13

-- 3. 069

-- 3.303

-- 3.355

-- 4. 576

-- 6. 398

-- 6. 756

-- 8. 002

-- 8.627

-- i0. 287

-- 12.826

-- 3. 296

-- 3.530

-- 3.582

-- 4. 634

-- 6. 619

-- 6. 983

-- 8. 229

-- 8.625

-- i0. 285

-- 12. 882

-- 0. 227

-- 0. 227

-- 0. 227

-- 0. O58

-- 0.221

-- 0. 227

-- 0. 227

0.0O2

0.002

-- 0. 057

S(6) 4.278 X 10 -35

so]- _ NaSO4

MgSO4

Kso7 CaSO4

FeS04

HSOs FeSO +

UO2SO4

USO 2+ _

Fe ( SO4 )2

uo2 ( so4 )~- U(SO4)2

4.110 • i0 -35

1.411X 10 -36

2.536 X i0 -3v

2.160 • -38

0.000

0.000

0.000

0.000

0.000

0.000

0.000

0.000

0.000

2.423 X 10 -35

1.245 X 10 -36

2.546 • 10 -37

1.906 X 10 -3s

0.000

0.000

0.000

0.000

0.000

0.000

0.000

0.000

0.000

- 34.386

- 35.850

-- 36.596

-- 37.666

- 40.300

-- 42.335

-- 43.506

- 65.480

-- 73.420

- 81.276

- 98.594

-- 106.334

- 111.832

- 34.616

- 35.905

- 36.594

-- 37.720

- 40.298

- 42.333

-- 43.562

-- 65.535

- 73.419

-- 81.503

-- 98.650

-- 106.560

- 111.830

- 0. 229

- 0.054

0.002

- 0.054

0.002

0.002

-- 0. 056

-- 0. 055

0.OO2

- 0. 227

- 0. 057

- 0. 227

0.002

U(3) 0.000 U 3+ 0.000 0.000 -- 48.795 -- 49.304 -- 0.510

U(4) 1.784 X 10 -11

U (OH) 5 1. 784 X 10 -11 1. 566 X 10 -1! -- i0. 748 -- i0. 805 -- 0. 057

U (OH) 4 7. 631 X 10 -18 7. 661 X i0 -18 -- 17. 117 -- 17. 116 0. 002

U (OH) + 3. 727 X 10 -25 3. 272 X 10 -25 -- 24. 429 -- 24. 485 -- 0. 057

U(OH) 2+ 2.976 X 10 -33 1.767 X 10 -33 -- 32.526 -- 32.753 -- 0.227

UOH 3+ 0. 000 0. 000 -- 41. 568 -- 42. 078 -- 0. 510

U 4+ 0. 000 0. 000 -- 51. 442 -- 52. 348 -- 0. 906

UCl 3+ 0. 000 0. 000 -- 53. 032 -- 53. 542 -- 0. 510

USO 2+ 0. 000 0. 000 -- 81. 276 -- 81. 503 -- 0. 227

U(S04)2 0.000 0.000 -- 111.832 -- 111.830 0.002 9+

U6(OH)15 0.000 0.000 -- 162.741 -- 167.328 -- 4.587

U(5) 1.477 X 10 -27

U02 + i. 477 X 10 -27 1.296 X 10 -27 - 26. 831 - 26. 887 - 0. 057

U(6) 2.232 X 10 -28

uo2 (c03)~- u02 (c03)s[- UO2CO3

UO2OH +

UO 2+

UO2CI +

(UO2)2 (OH) 2+

UO2SO4

(U02)3 (OH) +

uo2 ( so4 )~-

2.219 • i0 -2s

i. 246 X 10 -3o

5. 964 X 10 -35

1.893 X 10 -36

0.000

0.000

0.000

0.000

0.000

0.000 X i0 +~176176

2. 755 X10 -29

7. 396 X 10 -31

5. 988 X 10 -35

1.662 X 10 -36

0.000

0.000

0.000

0.000

0.000

0.000 XI0 +~176176

-- 27.654

-- 29.904

-- 34.224

-- 35.723

-- 41.285

-- 43.767

-- 66.552

-- 73.420

-- 85.367

-- 106.334

-- 28. 560

-- 30. 131

-- 34. 223

-- 35. 779

-- 41. 512

-- 43. 824

-- 66. 779

-- 73. 419

-- 85. 423

-- 106. 560

-- 0. 906

-- 0. 227

0.002

-- 0. 057

-- 0. 227

-- 0. 057

-- 0. 227

0.002

-- 0. 057

-- 0. 227


Saturation indices.

Phase SI log IAP log KT

Anhydrite - 37.97 - 8.95 29.02

Aragonite -- 2.47 -- 10.80 -- 8.34

Artinite 1.00 10.60 9.60

B_UO2(OH)2 -- 25.19 -- 16.86 8.33

Brucite 0.84 17.64 16.79

Calcite -- 2.33 -- 10.80 -- 8.47

CH4(g) 22.68 -- 17.42 -- 40.10

C02(g) -- 6.52 -- 24.68 -- 18.16

Dolomite -- 0.84 -- 17.84 - 17.00

Epsomite -- 36.71 -- 5.19 31.52

Fe -- 2.28 13.71 15.99

Fe (OH) 2.vClo.3 -- 3.04 6.95 9.99

Fe2 (S04)3 -- 177.11 -- 46.48 130.62

Fe3(OH)s -- 12.41 33.88 46.29

Ferrihydrite - 6.93 10.99 17.92

FeS(ppt) 0.25 -- 3.67 -- 3.92

Goethite - 2.54 10.99 13.53

Greigite - 9.41 - 28.39 - 18.97

Gummite - 30.05 - 16.86 13.19

Gypsum - 37.76 -- 8.95 28.81

H2(g) 1.77 1.81 0.04

Halite -- 6.10 -- 4.52 1.58

Hematite -- 0.07 21.98 22.06

Huntite -- 1.95 -- 31.92 -- 29.97

Hydromagnesite -- 1.76 -- 10.52 -- 8.77

Jarosite-H -- 106.99 -- 12.67 94.32

Jarosite-K -- 97.30 -- 5.69 91.62

Jarosite-Na -- 98.94 -- 3.73 95.22

Lepidocrocite -- 3.41 10.99 14.40

Lime -- 18.92 13.87 32.80

Mackinawite 0.98 -- 3.67 -- 4.65

Maghemite -- 10.47 21.98 32.45

Magnesite 0.99 -- 7.04 -- 8.03

Magnetite 4.08 33.88 29.80

Melanterite -- 42.12 -- 10.93 31.19

Mg-Ferrite -- 3.21 39.62 42.83

Mirabilite - 37.48 - 4.94 32.55

Natron -- 5.48 -- 6.79 -- 1.31

Nesquehonite -- 1.42 -- 7.04 -- 5.62

02(g) -- 86.75 -- 3.63 83.12

Periclase -- 3.87 17.64 21.51

Portlandite - 8.80 13.87 22.68

Pyrite - 2.57 - 21.05 - 18.48

Rutherfordine - 29.86 - 41.54 - 11.68

CaS04

CaC03

MgC03 :Mg (OH) 2:3H20

U02 (OH) 2

Mg ( OH ) 2

CaC03

CH4

C02

CaMg ( C03 ) 2

MgS04 : 7H20

Fe

Fe (OH) 2.vClo.3

Fe2 ( SO4 ) 3

Fe3 (OH) 8

Fe (OH) 3

FeS

FeOOH

Fe3S4

UO3

CaS04 : 2H20

H2

NaCI

Fe203

CaMg3 (C03) 4

Mg5 (C03) 4 (OH) 2 : 4H20

(H30) Fe3 (S04) 2 (OH) 6

KFe3 ( S04 ) 2 (OH) 6

NaFe3 (S04) 2 (OH) 6

FeOOH

CaO

FeS

Fe203

MgC03

Fe304

FeS04 : 7H20

MgFe204

Na2S04 : 10H20

Na2C03 : 10H20

MgC03 : 3H20

02

MgO

Ca(OH) 2

FeS2

U02C03

136 J. B 6 h m et al.

(continued)

Phase SI log lAP log KT

Schoepite

Siderite

Sulphur

Thenardite

Thermonatrite

U308 (C)

U409 (C)

UO2 (am)

UO3 ( C )

Uraninite

- - 25.05 -- 16.86 8.19

-- 2.23 -- 12.78 -- 10.55

-- 15.27 -- 17.38 -- 2.11

- - 38.41 -- 4.93 33.48

-- 6.91 -- 6.79 0.13

-- 50.58 -- 48.77 1.81

- - 32.90 -- 62.00 -- 29.11

- - 9.55 -- 15.05 -- 5.50

-- 27.37 -- 16.86 10.50

- - 3.92 -- 15.05 -- 11.13

u02 (OH) 2 : H20

FeC03

S

Na2S04

Na2C03 : H20

U308

U409

UO2

U03

U02

End of simulation.


Chapter 6

On-site column experiments

Mih~ily Cs6vfiri and Gfibor Simoncsics

A. Introduction

Laboratory batch and column experiments using artificial groundwater are usually easy to perform and control. Literature sources (Cantrell et al., 1995; Naftz et al., 1999; Morrison et al. 2OOl), and new results reported in Chapters 3 -5 of this volume, show that both elemental iron (Fe ~ also known as zero-valent iron, ZVI) and hydroxyapatite (HAP) are very promising reactive materials for use in permeable reactive barriers (PRBs) to remove uranium from contaminated groundwater. Nevertheless, purely laboratory-based experiments have the inherent disadvantage of fairly artificial boundary conditions, which might differ significantly from real field conditions for a number of crucial parameters and groundwater constituents, such as temperature, pH, dissolved oxygen content, redox conditions and the balance of the carbonate cycle.

To demonstrate that the aforementioned reactive materials are feasible for in situ

treatment of uranium-contaminated groundwater, on-site column experiments were designed and operated to overcome these problems. Two types of experiments were performed: small-scale in situ column experiments, where the columns filled with a reactive material were placed directly in the groundwater by installing them into monitoring wells, and large-scale column experiments located on-site (but ex situ) close to monitoring wells and using contaminated groundwater pumped from these wells for the experiments.

The experiments were performed on-site at a location where the installation of an experimental pilot-scale PRB using elemental iron (Chapter 1 O) was planned. The test site is located in a small valley downstream of a waste rock pile of the former uranium milling site near Prcs, Hungary (see site description in Chapter 9). The shallow groundwater aquifer in this valley has a direct connection to the drinking water aquifer of the city of Prcs. Objectives of the on-site column experiments were to prove the effectiveness of elemental iron and hydroxyapatite for removing uranium from groundwater under field conditions, and to quantify the changes in the groundwater quality caused by these materials. Therefore, these field column experiments were carried out prior to the planning and installation of the pilot PRB system.

The experiments were part of the decision process on the design of the planned PRB, especially on the question of whether a reactor-type installation (as in a funnel-and-gate system) or a continuous reactive barrier should be built. The flow rate through the reactive material in a reactor-type PRB is usually much higher than in a continuous reactive barrier.

138 M. Csrvrri, G. Simoncsics

Therefore, the on-site column tests had to provide data for both options. The small-scale in situ column experiments installed in monitoring wells had a relatively short residence time and represented the case of a reactor-type PRB, while the large-scale on-site column experiments, with their relatively long residence times, were a model of a continuous reactive barrier.

Materials used in the on-site column experiments were elemental iron in two forms (shredded cast iron and waste steel fibres from tyre recycling) and hydroxyapatite, as described in Chapters 3 and 4.

B. Columns in monitoring wells

l. Experimental set-up

The basic experimental approach was to perform column experiments directly in the groundwater by placing the reactive material into monitoring wells, thus ensuring that the experimental conditions, including temperature and redox conditions, were as close to real field circumstances as possible. These in situ experiments were carried out at a location downstream of waste rock pile III at the Prcs site. The column tests were performed in existing monitoring wells (Hb-0l/l and Pe-04). These monitoring wells are located at or close to the site of the planned experimental PRB (Chapter 10). Water from the wells has been used in previous laboratory column experiments. Only the performance of elemental iron was investigated in down-hole column tests. Two types of in situ columns have been designed for these tests (Figs 6.1 and 6.2): a column consisting of two subsections, each with different iron-to-sand ratios (first section: 10% elemental iron, second section: 50% elemental iron) and a single column with an iron content of 10% in sand (all contents given as percentage by volume). The columns were placed in the monitoring wells Hb-01/l and Pe-04 (Fig. 6.3) and operated based on the siphon principle. The water passed through the columns was discharged into a nearby ditch. Water was sucked through the columns by the hydraulic head difference between the water level in the wells and the ditch (Figs 6.1 and 6.2). Uranium concentrations in the groundwater were around 1000 Ixg/l.

2. Results

The in situ column tests started in October 2001 and lasted till April 2002. During wintertime the columns were operated only occasionally because the discharge was frozen. The data collected are listed in Table 6.1. The flow rates through the in situ columns were 7.6 X 10-5-1.1 x 10 -4 m/s, resulting in residence times for the groundwater within the columns of 1.22-4.21 h. More than 98% of the uranium was removed even during this short residence time, reducing uranium concentrations from around 1000 Ixg/l down to around 10-15 Ixg/l. The water chemistry of the column effluents was almost independent of the iron content (10-50%) at the flow rates investigated. The concentration of some groundwater constituents changed during the experiments, for example Ca and bicarbonate concentrations both showed a significant

On-site column experiments 139

Figure 6.1. Schematic diagram of the column experiment implemented in monitoring well Hb-O1/1.

decrease. The concentration of dissolved iron, which was undetectable in column influent, was around 20 mg/l in the column effluents (Table 6.1).

Iron concentration in the effluent is a function of the pH. Precipitation of iron in the form of siderite (FeCO3) occurs at pH values above 7.5, and this controlled the iron concentrations in the down-hole column experiments. The pH value increased with increasing residence time (Fig. 6.4).

C. Large-scale field column experiments

Experiments using in situ columns placed in monitoring wells were performed at high flow rates typical for funnel-and-gate systems, as compared to those of continuous barriers. To obtain data for lower flow rates, large diameter columns were built and operated on-site.

140 M. Cs6vdtri, G. Simoncsics

Figure 6.2. Schematic diagram of the column experiment implemented in monitoring well Pe-04.

The tests were performed on both elemental iron (shredded cast iron and waste steel fibres from tyre recycling) and hydroxyapatite. The tests were also aimed at studying the effect of elemental iron on the groundwater composition, especially on the concentration of dissolved iron in the treated water of the column effluents at the flow rates expected under field conditions.

1. Experimental set-up

Five columns were constructed and filled with reactive materials. The columns filled with elemental iron and sand mixtures consisted of two sections: the first section contained approximately 10-15% (by volume) of iron particles, while the second one had a higher

Figure 6.3. Columns being prepared and placed into the monitoring well Hb-Ol/1.

Table 6. I. Surnrnarised data for down-hole columns in the monitoring wells.

Time period Flow rate Residence time pH U Specific electric TDS E F e -~ ~- ~- ~

W S ) (h) (pg/l) conductivity (mgll) ( m a ) (pS/cm)

Well H t - O M October 2001-March 2002

Well Pe-04 October 2001 -April 2002

T- 1 October 2001 November 2001 December 2001 January 2002 February 2002 March 2002

Average values Change in composition

T-2 October 2001 November 2001 December 200 I January 2002 February 2002 March 2002

Average values Change in composition

8.86 X

Frozen Frozen 7.28 X lo-'

1.01 x

8.65 x 1 0 - ~

8.72 x 1 0 P

1.64 x 10 -4

1.01 x Frozen Frozen

9.07 X 10 -' 1.1ox 10

8.55 x

1.27 1.10

2.20 1.30

1.47

0.68 1.70

1.30 1.20

1.22

7.36

7.46

7.57 7.49

7.58 7.58

7.56 0.20

7.33 7.44

7.35 7.53

7.41 0

828

1030

5.5 19.5

22 15

16 -813

6.5 14

9 8

9 - 819

1714

1746

1713 1532

1560 1386

1548 - 166

1569 1504

1521 1346

1485 - 229

1220 <0.1

1185 10 .1

1199 23 1026 25

1092 12 929 12

1062 18 -158 18

1098 37 1008 22

1065 28 902 14

1018 25 -202 25

188

173

188 133

121 102

136 - 52

175 117

112 82

122 - 67

62

63

62 61

61 59

61 - I

62 62

59 60

61 - 1

648

610

597 640

560 505

576 - 73

592 548

520 479

535 - 113

Table 6. I . (continued) e

R ___- ~ ~ ~ - _ _ _ -~ --

Time period Flow rate Residence time pH U

W S ) (h) ( & g o

Specific electric TDS XFe Ca Mg HC03 conductivity (mgfl) (mg/l) (mg/l) (rngfl) (mg/l) ( u.S/cm)

Pe-04 October 2001 November 2001 December 2001 January 2002 February 2002 March 2002

I .23 X lo-" 2.86 7.59 13

Frozen Frozen

5.71 x lo-' 4.94 7.73 12

5.71 X 10-' 4.94 7.78 20 6.76X lo-' 4.11 7.82 12

1617 1099 21 127 60 609 1606 1110 15 122 65 524

1482 1010 13 41 65 402 1420 1050 9 88 63 505

Average values 7.62 X 4.21 7.73 14 1531 1067 15 95 63 510 Change in composition 0.27 -1016 -215 -118 15 -79 0 - 100

- -~ ~~


7.9

7.8

7.7

...r 7.6

7.5

7.4

7.3

J f

O o f �9

o 6 Residence time (hours)

Figure 6.4. pH as a function of groundwater residence time in the down-hole column tests.

iron concentration. The purpose of this set-up was to remove the influent dissolved oxygen in the first section of the column. It was expected that the permeability would remain almost unaffected in this section because the corrosion products would be dispersed in a larger volume of sand due to the lower iron concentration. The pollutants could then be removed in the second section. Hydroxyapatite was used in a 20% HAP to 80% sand mixture (by volume).

The columns had a height of 180 cm and an internal diameter of 15 cm. The construction of the columns is illustrated in Figure 6.5. The experimental parameters for these tests are compiled in Table 6.2.

The columns were placed in supporting tubes (d = 200 mm) into a 2 m deep open shaft. The hydraulic gradient between the monitoring well (Pe-O4) which provided the uranium- contaminated groundwater and the location of the column experiments was high enough to provide an appropriate flow rate in the columns (Fig. 6.6). The ends of the columns were equipped with ports for inflow and outflow as well as with sampling points. The outflow from the columns overflowed into the shaft, and thence into the drainage ditch. Photographs of the installation are presented in Figure 6.7. During the experiments hydrogen gas formation was observed in the columns that contained elemental iron particles and this affected the flow rate. Therefore, gas-collecting vessels were mounted on the top of these columns (see Fig. 6.7).

2. Results

The column experiments lasted for more than 6 months. Uranium contaminated groundwater was fed into the on-site columns from monitoring well Pe-O4, which was at a distance of approximately 35 m from the columns. Water was distributed and adjusted to the desired flow rate by valves. However, stability of the flow rate in the columns operated with elemental iron was influenced by gas formation. Stable flow rates were observed in the column with HAP where no gas formation was detected. The flow rates were calculated from the volume of water passing through the columns. The flow rate was kept at about 10 -6 trots, which is comparable with the estimated groundwater flow rate at the site. The residence time of the water in the columns was between lOO and 250 h.

144 M. Csb'vfri, G. Simoncsics

Figure 6.5. Schematic diagram of the large-scale field columns.

A variety of data were collected during the experiments: flow rate (almost daily), pH, specific electric conductivity, total dissolved solids (TDS), redox potential (Eh), U, Ca, Mg, Fe and C 0 2 - / H C 0 3 concentrations. The results are summarised in Table 6.3. Pore water in the columns was sampled from time to time at different heights of the column to retrieve spatial information on the distribution of selected groundwater constituents. Concentration profiles along the columns were measured, and the results are listed in Table 6.4.

Uranium was removed from the contaminated groundwater in all columns with a high degree of efficiency. The concentration in the effluent dropped to or even below lO Ixg/l

Table 6.2. Experimental parameters for the field column experiments (all percentages by volume).

FLC l FLC 2 FLC 3 FLC 4 FLC 5

1. Section: 0.5 m, l. Section: 0.5 m, l. Section: 0.5 m, 13% iron, 13% iron, 13% iron, 87% sand 87% sand 87% sand

2. Section: 1.0 m, 2. Section: 1.0 m, 2. Section: 1.0 m, 25% iron 50% iron 100% iron (0.2-3 mm), (0.2-3 mm), (l - 3 mm) 75% sand 50% sand

10% waste 20% HAP, steel fibres, 80% sand 90% sand


Figure 6.6. General layout plan of the test site (schematic, not to scale).

Figure 6. 7. On-site column experiments; note the gas collectors on the columns filled with elemental iron.

Table 6.3. Summary of results from the field columns

Column Date Flow rate Residence pH U Specific TDS ZFe Ca Mg C1 Sod ~

(dS) time (h) (Kgll) electncal ( m g ) (rng/l) (rng/l) (mgll) (mg/l) (mg/l)

Feed well Pe-04, average

FLC 1, Fe 25% 05/2002 06/2002 07/2002 1 012002 1 1 I2002

Average

FLC 2, Fe 50%' 0412002 0512002 0612002 0712002 I012002 1112002

Average

FLC 3, Fe 100% 05/2002 06/2002 07/2002 10/2002 1112002

Average

07/2002 1012002 11/2002

Average

FLC 4, Fe fibres 100%

9 . 7 ~ 19s

7 . 9 ~ 238 2.2x 10-6 85 3.5x 10 55

1 . 4 ~ lo-" 122

4 . 7 ~ lo-' 40

1.2X 156

1.2X 154 1.3X 10 143 8.9X lo-' 213

2.9X 10 ' 66 3 .5x 1 0 - 6 55

2 . 4 ~ 1 0 ' 112

1 . 8 ~ lo-' 106

7.7 X lo-' 246 1.1 x 10 ' 173 3.5 x 1 0 - 6 55

9 . 7 ~ lo-' 19s

1 . 6 ~ 155

1 . 0 ~ 182 1.1 X lo-' 173 3.8X lo-' 50

2 . 0 ~ 10 135

7.22 1143 1671

9.61 848 8.47 10 925 8.90 (10 1071 8.80 < I 0 1279 9.80 1 1 0 827

7.60 (10 825

8.69 1150 8.46 3 1079 9.20 9.6 1001 8.92 5 903 8.88 5 1028 9.40 5 1164

8.93 5 1054

8.81 20 980 8.55 5 978 8.31 7 832 9.81 1103 9.36 1111

8.97 6 1001

7.93 37 8.57 10 1212 8.00 7.5 1492

8.17 18 901

1197 (0.1 166

620 <0.5 33 605 <0.5 63 725 c0.5 19 822 1 0 . 5 30 618 cO.5 46

565 10.5 32

850 <0.5 56 700 <o.s 44 624 <0.5 46 560 10.5 23 620 10.5 78 825 CO.5 53

697 1 0 . 5 50

680 c0.5 37 690 10.5 54 603 1 0 . 5 42 706 <0.5 33 740 10.5 48

684 (0 .5 43

771 1 0 . 5 62 807 10 .5 39

1082 3.1 95

887 nd 65

63

11 10 10 56 43

22

32 33 27 21 38 64

36

14 15 15 42 49

27

27 36 63

42

-

43

43

43

43

46 43 43

43

43

43 43

43

43 43 43

43

--

38 1

290

250 239

260

32 1

32 1

312

304

308

39s

354

250

-

nd

nd 6 6

45 50

18

12 11 12 6

79 81

34

38 24 12 84 33

38

nd 15 6

7

~- 591

121 158 202 270 254

168

23 1 228 191 199 25 8 234

224

158 176

142 263 <OP

a,

162 % 9

374 a

3 443 n

180 ~ 2.

534 420 $ 4.

Table 6.3. (continued)

Column Date

F'LC 5, HAP 20% 05i2002 06/2002 07i2002 I0/2002 11/2002

~~

1.3 X lo-"

5.5 x LO l.OX10 1.8 x 10-6

1.1 x 10-6

Residence pH U Specific TDS ZFe Ca Mg time (h) ( I J , ~ ) electrical (mgA) (rngA) (mgh) (rng/l)

conductivity i uSicrn)

141 6.91 9 1138 743 nd 177 7.35 6 1258 838 nd 104 40 343 7.98 1271 900 nd 100 44 189 7.60 9 1372 962 nd 98 57 108 7.45 7 1634 1460 nd 127 77

43 43 43 43 43

Average 1.2 X 170 7.46 6 1335 981 nd 86 44 43 nd 362 ~~~~ - -

nd = not detected.

Table 6.4. Concentration profiles along the field columns (temporal average).

Height pH U Specific TDS C03 HCO, Ca Mg so4 CI ZFe Redox (cm) (pgf l ) conductivity (mg/l) (mgfl) ( m u ) (mg/l) ( m d ) ( m d ) (mg/l) (mg/l) potential

-

(Wcm) (mv)

c P 00

25% Fe/75% sand 150 8.70 < I 0 125 9.00 6 I00 8.86 7 75 8.27 4 50 7.98 9

50% Fe/50% sand 150 8.67 < 10 125 9.46 7 100 9.17 5 75 8.52 10 50 8.00 17

100% Fe (coarse) 150 9.12 <10 125 9.91 7 100 9.70 4 75 9.42 2 50 7.94 10

100% Fe (waste steel fibres) 170 8.00 <10 150 8.42 12 125 100 7.90 10 75 50 7.7 1 15

953 1020 1127 i217 1402

1033 1050 1113 1233 1328

994 955

1047 1097 1307

1212 1294

1455

1455

669 740 770 880

1230

648 744 744 824 984

674 590 610 660 880

807 880

940

990

26 239 26 225 26 266 4 335

nd 388

26 205 35 228 22 205 6 272

405

32 182 42 125 36 133 24 170 nd 396

438 nd 439

nd 414

nd 48 1

34 44 53 59

100

35 40 40 45 84

31 26 33 29 76

76 80

90

92

26 36 45 50 50

34 34 49 47 46

29 28 34 41 55

44 62

57

63

227 279

362 436

329

329

378

268 330 395 360 378

41 8 403

41 1

403

43 43 43 43 43

43 43 43

43 43

43 43

<0.5 1.15 2.44 4.9 5.68

<0.5 3.39 3.3 6.88 9.32

< 0.5 0.28 0.86 1.15 4.7

3.1 15.52

1 I .68

4.62

- 93 - 104 - 81 - 81

- 283 - 178 - 115 - 12

- 242 - 248 - 206 -179 5

-17 9 ?? 2.

-170 9

9,

-105 8 3

- 110 5.

9 Table 6.4. (continued)

20% HAP/80% sand 150 7.27 < I 0 135 7.03 11 110 7.01 7 85 7.14 5 60 7.22 15 35 7.27 6 10 7.29 17

Specific conductivity bS/cm)

1296 1300 1315 1366 1413 1400 1421

991 1032 1032 1128 1128 1160 1176

nd 439 404 414 469 509 515 539

100 105 107 119 129 131 142

45 55 57 58 57 55 55

Not measured Not measured Not measured Not measured Not measured Not measured Not measured

-, -. C1 Z F e Redox 3 (mg/I) (mg/l) potential c?

w 43 Not measured B 43 2.

43 43 6 43 43 43

(mV) $

nd = not detected. Reference electrode for measurement of redox potential: saturated calomel electrode.

150 M. Csrvdri, G. Simoncsics

from the initial value of 1100-1200 txg/l. Uranium removal was somewhat less efficient in the columns with waste steel fibres than in the other columns. However, even here the concentration decreased to the level of 10-20 Ixg/l. The concentration profiles (Table 6.4) show that the uranium was removed in the first sections of the columns.

The water chemistry was changed by all the columns containing elemental iron. The TDS concentration decreased, most likely due to precipitation of calcium and magnesium carbonates. TDS dropped in the columns operated with elemental iron on average from the initial 1197 mg/l to approximately 650 mg/l in the column effluents, almost independently from the volumetric concentration of iron particles in the iron/sand mixtures. At longer residence times TDS dropped to even below 600 mg/l. This means that 500 mg of precipitates per litre of treated water were being formed. Sulphate concentrations also decreased to some extent, most probably due to reduction to sulphide. The pH value increased during the treatment with elemental iron from approximately 7.2 to 9. Hydroxyapatite also proved to be effective for uranium removal from the contaminated groundwater. However, the groundwater chemistry underwent only minor changes compared to iron as a reactive material.

Dissolved iron was detected only in the effluent of the column containing waste steel fibres, and then only at low residence times (< 50 h). It can be expected that the iron concentration in the effluent of an elemental iron barrier at the site in P6cs will remain below 1 mg/l under natural field conditions. However, it is not clear whether or not the results of the column experiments were to some extent influenced by oxygen, which might have entered the columns at the outflow ports.

D. Conclusions

On-site column experiments at the former uranium ore mining and processing site near Prcs, Hungary, have proved that uranium can be efficiently removed from the local groundwater using elemental iron and hydroxyapatite as reactive materials, decreasing uranium concentrations from an initial approximately 1000 Ixg/l down to < 10-20 Ixg/l. Both elemental iron types tested (shredded cast iron and waste steel fibres from tyre recycling) proved to be feasible for the treatment of the uranium-contaminated groundwater at Prcs. A relatively low residence time of 1-5 h was sufficient for uranium removal from the groundwater.

To avoid a permeability reduction in an elemental iron barrier, the reactive matrix has to be mixed with sand. The experiments have shown that even waste steel wire at low volumetric concentration (--~ 10%) is effective for uranium removal.

The composition of the treated groundwater undergoes significant changes during its passage through an elemental iron column: pH increases, while TDS and Ca, Mg, bicarbonate and sulphate concentrations decrease. These changes are unfavourable from the point of view of the hydraulic performance of a PRB, because they can lead to a significant decrease in hydraulic conductivity. The bulk of the precipitates were formed in the first section of the columns. When hydroxyapatite is used as a reactive material, the chemistry of the treated groundwater undergoes less dramatic changes than with iron, so the effect caused by precipitation of secondary minerals is expected to be less of a problem.


Further it was shown that the outlet composition of the groundwater, percolated through the elemental iron columns, depended significantly on the residence time, especially for dissolved iron. Elevated concentrations of dissolved iron in the range of 15-25 mg/l only occurred in the effluent of elemental iron columns when the flow rate was high, as shown in the in situ column experiments with residence times around 1 -5 h. At low flow rates, iron concentrations remained below < 0.5 mg/l.

As an overall conclusion it can be stated that both elemental iron and hydroxyapatite can be effective in removing uranium under in situ conditions, and the choice between these materials should be based on site-specific conditions.

Acknowledgement

The authors sincerely thank Jtilia Grit and Benc6n6 l~va Tdth for their excellent analytical work.

References

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.

Morrison, S.J., Metzler, D.R., Carpenter, C.E., 2001. Uranium precipitation in a permeable reactive barrier by progressive irreversible dissolution of zerovalent iron. Environ. Sci. Technol. 35 (2), 385-390.

Naftz, D.L., Davis, J.A., Fuller, C.C., Morrison, S.J., Freethey, G.W., Feltcorn, E.M., Wilhelm, R.G., Piana, M.J., Joye, J., Rowland, R.C., 1999. Field demonstration of permeable reactive barriers to control radionuclide and trace element contamination in groundwater from abandoned mine lands. In: Morganwalp, D.W., Buxton, H.T. (Eds), U.S. Geological Survey Toxic Substances Hydrology Program - Proceedings of the Technical Meeting, Charleston, South Carolina, March 8-12, 1999: U.S. Geological Survey Water-Resources Investigations Report 99-4018A, p. 281-288.


Chapter 7

New barrier materials: the use of tailored ligand systems for the removal of metals from

groundwater

D.I. Stewart, David E. Bryant, Catherine S. Barton, Katherine Morris and Mih4ly Cs6v~iri

A. Introduction

A research project was undertaken collaboratively at the University of Leeds to develop and test new ligand-based active materials for treating uranium-contaminated groundwater. These materials have great potential for use in permeable reactive barriers (PRBs) because they can be designed to have good selectivity for the uranyl cation (or for other heavy metals). This means that sequestering other species in solution will not use up their capacity, and that they will have minimal other impact on the groundwater chemistry. Also, by selecting a suitable support material, they can be made robust and durable enough for engineering use, with a particle size to meet the permeability requirements of an engineered barrier. However, if such materials are to be used in practice, they must also be cost-effective.

This chapter starts by outlining the approach taken to develop new ligand-based active materials for sequestering uranium from solution. It describes the synthesis of one such material, called PANSIL, and discusses the mechanism by which PANSIL sequesters uranium from solution. It goes on to describe an extensive test programme undertaken to determine the chemical conditions under which PANSIL will sequester uranium. It also reports continuous flow column tests, using both contaminated synthetic groundwater and contaminated groundwater from a uranium mine tailings disposal site in Hungary, designed to replicate flow and exposure conditions within a PRB. It then compares the performance of the newly developed active material with that of two commercially available ion-exchange resins. It concludes by comparing data on the cost effectiveness of PANSIL, the ion exchange resins, hydroxyapatite, and iron oxide pellets in treating uranium-contaminated groundwater.

In the synthesis of the active materials a general strategy was adopted of attaching the uranium sequestering ligands to a silica surface. Silica (in the form of sand) was selected as the support because the surface is easily modified to provide binding sites for attaching the new active matrices, it has high mechanical strength, and is available in a range of particle sizes, which allows barrier permeability to be controlled. Therefore, the main challenges of the synthesis were to find ligands selective for uranium, and to develop methods for attaching these ligands to silica. Subsequently, the challenge was to develop

154 D.L Stewart et al.

a test programme that not only measured the performance of PANSIL with contaminated groundwater under typical PRB conditions, but also elucidated the factors controlling its performance and thus understand the effect of minor changes in groundwater chemistry on its efficacy. Geochemical modelling of the groundwater and other test solutions to determine the major uranium species under given test conditions were essential to achieve the latter objective.

B. Concept and development

Uranium exists in aqueous solution as the very stable UO 2+ (uranyl) cation, which has a linear shape. The ligand systems which are the most selective for uranium are those with coordinating atoms of oxygen or nitrogen arranged in a planar ring surrounding the uranyl unit orthogonal to the O=U=O axis. Three types of these organic ligands were identified:

1. based on calixarenes; 2. based on salens; 3. based on polymers.

Within the project, attempts were made to synthesise all three types of active material with varying degree of success. The aim of this section is to review the attempts made to develop appropriate active materials, highlighting the pitfalls associated with different approaches and describing the development of a new active material that has been named PANSIL.

1. Calixarenes

It has been reported that five-, six- and eight-member ring calixarenes will remove uranyl from aqueous solution with a very high degree of selectivity (Kabay et al., 1993). Thus, the possibility of using calixarenes as ligands in the active material was investigated. Calix[8]arene was successfully synthesised and was shown to effectively remove uranyl from aqueous solution. However, in order to attach calixarene molecules to a silica surface, in accordance with the general strategy, the calixarene has to be chemically modified. This proved to be a technically feasible but inefficient process. The approach using calixarenes was ruled out on economic grounds and not pursued (see Stewart et al. (2003), for more details).

2. Salens

Salens represent another promising multidentate ligand system that is believed to complex uranyl cations. Salens are formed from one molecule of a di-amine with two molecules of salicylaldehyde, one for each amine group. To attach salens to silica, one of these two fragments needs to be suitably modified. Many different approaches were attempted, but as with calixarenes this modification proved to be impractical (see Stewart et al. (2003)).

New barrier materials 155

3. Polymers

The polymer selected was polyacryloamidoxime resin, which is a modified form of commercially available polyacrylonitrile (PAN); see Scheme 1. Modified PAN-based fibres have been used for complexing a range of pollutant metals, including the removal of uranium from seawater (Vernon and Shah, 1983; McComb and Gesser, 1997; Suzuki et al., 2000). Indeed it is apparent that modified PAN resin can show a degree of selectivity for heavy metals, for example preferentially removing gold in the presence of first row transition metals (Lin et al., 1993b). Modified PAN-based fibres immersed in seawater have been shown to preferentially adsorb uranium, vanadium, cobalt, nickel, iron, manganese, in the presence of similar concentrations of copper and zinc, and far larger concentrations of calcium and magnesium (Potts et al., 1990). In another study on modified PAN resin, the resin showed great affinity for U(VI), some affinity for Pb(II) and Cu(II), but less affinity for Cd(II), Cr(III), Hg(II) and Zn(II) (Rivas et al., 2000). However, even though modified PAN fibres demonstrated selectivity for contaminants including uranium, they were considered unsuitable for use in PRBs due to their reduced mechanical strength after reaction, low water permeability, and the engineering difficulties associated with achieving a particular groundwater residence time with a fibrous material (Lin et al., 1993a). The approach taken here was to react a powdered form of PAN with hydroxylamine hydrochloride to form the active resin and then to deposit this resin from solution onto the surface of quartz sand to form a thin film coating. Thus, the overall mechanical properties of the sand were retained.

NH2OH

Scheme 7.1. Reaction of PAN to polyacryloamidoxime.

/__NHI N

I OH

Modified PAN-coated silica sand was successfully produced. The product has been named PANSIL. The synthetic route (described in Section C) uses readily available raw materials and is very straightforward with a high yield. Thus, PANSIL is a relatively low- cost active material on an appropriate matrix for PRB engineering.

C. The preparation of PANSIL

The synthesis stage of the project produced one relatively low-cost product in an abundant yield a modified PAN-coated silica sand, PANSIL. The synthesis of PANSIL is described in detail below:

Polyacrylonitrile powder (0.1 mol) was suspended in a solution of hydroxylamine hydrochloride (6.9 g, 0.1 mol) in 125 ml methanol. Sodium hydroxide (4.0 g, 0.1 mol) was added and the mixture refluxed for 4 h. The mixture was allowed to cool and then the


Figure 7.1. Electron micrograph of PANSIL.

polymer was filtered off and washed with deionised water and acetone and allowed to dry. The reaction was monitored by infrared spectroscopy and as synthesis progressed, the band due to the nitrile stretch at 2244 cm-1 diminished and the band due to the formation of an oxime at 1660 cm-1 increased. Not all functional groups were reacted after 4 h, but the reaction was stopped at this stage since it is believed that further modification has a detrimental effect on the coating properties of the modified polymer.

The modified PAN (polyacryloamidoxime) (2 g) was disaggregated by gently breaking-up the coarser particles using a mortar and pestle and dried in a desiccator, before dissolving in 250 ml dimethylsulphoxide (DMSO) at room temperature. At the same time 100 g of 0.6-1.2 mm diameter sand (cleaned with three 24-h washes with aqua regia leach and reconditioned with deionised water) was placed in a round-bottomed flask and heated under vacuum to 180~ for 2 h. The sand was allowed to cool and the vacuum in the flask was replaced with nitrogen. When the glassware was sufficiently cool to handle the DMSO, polymer solution was poured over the sand. The flask was evacuated to degas the solution and heated to reflux under vacuum (l 10~ for l 0 min then placed on a rotary evaporator to remove the DMSO, whilst ensuring gentle rotation for even coating of the sand. When the removal of solvent was complete and the mixture had a viscous fluid consistency, the coated sand was poured into a wide dish and allowed to stand for more than 4 days. During this time the polymer formed a gel and periodically relatively clean solvent was decanted off before stirring the mixture and repeating the procedure. Final drying was performed in a vacuum oven. An electron micrograph of PANSIL is shown in Figure 7.1.

D. Efficiency of contaminant attenuation

Following the successful synthesis of the PANSIL, it was necessary to assess the binding mechanism to PANSIL as well as evaluate the efficacy of the new material in scavenging


UO 2+ over a wide range of chemical conditions. The results of these experiments are described in this section.

1. Mechanism of contaminant removal by PANSIL

1.1. Column breakthrough test

Initially, simple column tests, using a rapid but approximate method of uranium analysis (Yong et al., 1996), were conducted to demonstrate that PANSIL was effective at removing uranium from solution and to give a rough estimate of PANSIL capacity. A sample of PANSIL (3.0 g), prepared as above, was mixed with fused alumina (12 g) as an inert diluent and placed in a column. The column was challenged with 1O-ml aliquots of uranyl nitrate solution (lOppm) which were collected as 1O-ml samples, keeping the column continuously immersed. Aliquots of the eluant were analysed for UO22+ and showed a total removal of uranium from aqueous solution of around 880 mg UO2+/kg of PANSIL.

1.2. Surface analysis

In order to examine the coating of PANSIL on silica, and the distribution of uranium in PANSIL samples after exposure to uranium, three samples were specially prepared for electron microscopy. Commercially supplied silica (particle size of 2 -5 mm, designed for chromatography) was used to support the modified PAN instead of sand, in order to maintain purity and consistency. The samples were

1. vacuum-dried silica exposed to uranium, 2. polyacryloamidoxime-coated silica, and 3. polyacryloamidoxime-coated silica exposed to uranium.

Sample (2) was primarily for visual study, whereas samples (l) and (3) were analysed for uranium using an energy dispersive X-ray (EDX) technique.

The silica (as supplied) was heated under vacuum to 180~ for 2 h to dry the surface. One portion of this dried silica (0.2 g) was suspended in a solution of uranyl nitrate in diethyl ether (0.2 mM) overnight. The solution was decanted off and the silica washed with fresh diethyl ether and allowed to dry in air to form sample (l). Polyacryloamidoxime was dissolved in DMSO, and this solution evaporated to dryness on a second portion of the dried silica using a rotary evaporator. The resultant solid was washed with ether and dried under vacuum as sample (2). A sub-sample of sample (2) was suspended in the uranyl ether solution as above to form sample (3). All samples were carbon coated for microscopy.

Analysis of sample (l) using energy EDX analysis detected no significant signal for uranium indicating that uranium does not sorb to silica in amounts detectable by EDX (a spectrum is shown in Figure 7.2). However, significant amounts of uranium were detected in EDX analysis of sample (3) (Fig. 7.3). The backscattered electron micrograph of sample (3) (Fig. 7.4) showed that uranium adhered to discrete locations on the coated silica (elements with a high atomic number show up as bright areas in backscatter images). Element mapping for silicon and uranium showed that uranium was detected only in those

cps

0

300--

200--"

100--"

0

eps

CI S

�9 _ m _ _ _

, m . . . . I . . . . ' r ' ' l . . . . m . . . . l . . . . ' . . . . l " - ' " " , - i ' , �9

2 4 6 8 E n e r g y O,ce" ,~

Figure 7.2. EDX spectrum of sample 1.

O

S i U


2 0 0 -

1 5 0 -

1 0 0 -

50--

O , , ' i . . . . I . . . . i . . . . I . . . . i . . . . I . . . . i . . . . I . . . . i ' ' '

2 4 6 8 E n e r g y ( k e " , O

Figure 7.3. EDX spectrum of sample 3.

areas where the silicon signal was masked (Fig. 7.5). Therefore, presumably the uranium was retained only in areas where the silica surface was coated with modified PAN (it should be noted that the coating on specimens prepared for surface analysis was less continuous than that on subsequent batches of PANSIL).

1.3. Infrared study

To assess the bonding mechanism of UO 2+ with the modified PAN, a series of samples were prepared for IR analysis: (i) A KBr disc was made of uranyl nitrate and its IR


Figure 7.4. Backscattered electron picture of sample 3 with uranium showing as bright areas on two silica particles.

spectrum recorded; (ii) uranyl nitrate (217 mg, 0.55 mmol) was dissolved in 10 ml DMSO, and its IR spectrum was recorded between NaCI plates; and (iii) polyacryloamidoxime (40 mg, 0.55 mmol) was dissolved in DMSO (10 ml) and exposed to uranyl nitrate solution. On mixing the polymer and uranyl solutions, the solution took on a yellow colour and on addition of water the polymer came out of solution. The polymer was washed with water and acetone, and its IR spectrum in KBr recorded.

The literature value of the asymmetric uranyl stretching frequency, Vouo, is quoted at 925 cm-1 (Potts et al., 1990). The asymmetric stretching frequency measured for UO 2+ in the KBr disc was 949 cm -1. The asymmetric stretching frequency measured for the combined modified PAN-uranyl nitrate solution in DMSO between NaCI plates was 954 cm-1. Finally, the polymer that had been exposed to uranyl nitrate exhibited a band due to the asymmetric stretch at 912 cm -1. Since the ~'ouo frequency decreases as the uranyl-ligand bond strength increases (McCabe et al., 1987), the shift in the asymmetric stretching frequency from 949 to 912 cm-1 indicates that the uranium coordinates more

Figure 7.5. Two further pictures of the same view of sample 3: (i) silicon X-ray emission (left), and (ii) uranium X-ray emission.


O

N ~ J ~ O

H

Figure 7.6. Schematic diagram showing uranyl sequestration by amidoxime.

strongly to modified PAN than with nitrate, and this effect cannot be attributed to residual DMSO left in the PAN.

1.4. Discussion of the mechanism of contaminant removal by PANSIL

Rivas et al. (2000) demonstrated the polyacryloamidoxime resin formed stable complexes uranyl ions by involvement of the amidoxime groups in the interaction with the uranyl ions. It is thought that the mode of interaction is the donation of lone pairs of electrons by the amino nitrogen and oxime oxygen to the positive metal centre to form a five-member ring (including the metal) in the plane orthogonal to the O=U=O axis (see Fig. 7.6); such tings are noted for their stability due to the minimum of strain in the bonds. The oxime oxygen can also undergo metal-assisted deprotonation further increasing the stability of the ring at circumneutral pH (Rivas et al., 2000).

2. Efficiency and durability of PANSIL

PANSIL was subjected to a series of batch and flow-through column experiments to determine the range of pH and solution chemistry over which it effectively sequesters uranium. In particular, the effectiveness of PANSIL at sequestering uranium from solutions typical of groundwater was investigated. The stability of PANSIL and its likely impact on groundwater chemistry were also evaluated.

With the exception of the column test with groundwater from a uranium contaminated site in Hungary, Ptcs, which was conducted separately (with details described below), all the tests reported below were conducted in the same laboratory, and all solution samples treated following the same protocol. The tests were conducted at room temperature (typically 22-25~ and no attempts were made to exclude atmospheric CO2(g). Solution samples taken for analysis were filtered (< 0.45 I~m), had their pH measured, then they were acidified by adding a few drops of concentrated nitric acid and stored at 4~ Their uranium concentration was measured using a spectrophotometric method based upon the complexing reagent 2-(5-bromo-2-pyridylazo)-5-diethylaminophenol (known as bromo-PADAP) which has a detection limit of --~0.2 mg/l (Johnson and Florence, 1971).

Prior to the main test programme, a series of increasing duration batch exposure tests were conducted on PANSIL with a uranyl nitrate solution at a liquid to solid (L:S) ratio of 30:1 (see Stewart et al. (2003), for details). These tests indicated that an exposure time of less than 4 h was sufficient for PANSIL to reach equilibrium with a uranyl nitrate solution


containing 10 mg of UO2+/I at ---pH 6, and routinely experiments were run with a 24 h equilibration time.

2.1. Determination of the sorption isotherm for uranium on PANSIL

PANSIL was added to stoppered glass bottles containing a uranyl nitrate solution whose pH was adjusted to a value between 4 and 5. The liquid-to-solid ratio was 30: l, and the uranium concentration was between l and 1000 mg/l. Control tests were conducted using acid washed sand instead of PANSIL. The bottles were shaken end-over-end for 24 h at room temperature (typically 22-25~ and the solution was sampled for analysis for UO 2+.

A second series of tests similar to those just described were conducted where the solutions contained equal molarities of UO 2+ and Pb 2+ at uranium concentrations between 1 and 1000 mg/1. The aim of these tests was to determine the sorption behaviour of uranium on PANSIL in the presence of a competing divalent heavy metal ion.

2.2. pH-controlled batch exposure tests

The effectiveness of PANSIL at extracting uranium from water over a range of pH was investigated by batch exposure tests. PANSIL was added to stoppered glass bottles containing a uranyl nitrate solution whose pH was adjusted to values between 1.5 and 9 using either HNO3 or NaOH (coveting the usual groundwater pH range of 5 - 9 and also that encountered in acid mine drainage). The uranium concentration was either l or 10 mg/1, and the liquid-to-solid ratio was 30:1. The bottles were shaken end-over-end for 24 h at room temperature and the solution was sampled for analysis.

The exposed PANSIL from four batch exposure tests at pH 5 - 6 (two at each uranium concentration) was subjected to two further leaching steps to examine the retention of UO 2+ on the PANSIL. Once the initial solution had been decanted off, the exposed PANSIL was twice shaken end-over-end in distilled water at a liquid-to-solid ratio of 30: l for 24 h. The water from both these steps was sampled for analysis.

Four longer duration batch exposure tests were conducted over the pH range of 8 -1 l with a solution containing 10 mg/l UO22+. In these tests, small samples of the test solution were taken for analysis after 24, 48 and 96 h. After 96 h the contaminant solution was decanted off, and the exposed PANSIL was subjected to three further leaching steps. In these leaching steps PANSIL was shaken end-over-end in distilled water at a liquid-to-solid ratio of 30:1 for 24 h. The water from each step was sampled for analysis.

2.3. Batch exposure tests with synthetic groundwater

Further batch exposure tests with a duration of 24 h were conducted on PANSIL using a solution representative of a natural groundwater contaminated with uranium. This "synthetic groundwater" was made up from laboratory grade reagents (200 mg/l CaCO3, 272 mg/1 CaSO4, 194 mg/l 4MgCO3.Mg(OH2).5H20, 252 mg/l NaHCO3, 75 mg/l KCI), and the pH was adjusted to 7 using H2S04 (0.6 ml/l of 0.1 mol/l). After pH adjustment an undissolved residue remained, so the synthetic groundwater was filtered and its


Table 7.1. Measured composition of the synthetic groundwater prior to addition of UO2(NO3)2.

Na + K + Ca 2+ Mg 2+ C1- SO 2- HCO3

Concentration (mg/l) 64.9 38.4 31.4 43.0 35.5 252.5 181.5

Cations were measured by ICP-OES, anions were measured by HPLC-IC except for carbonate, which was measured by the flow injection method (Hall and Aller, 1992). The error in charge balance was less that 2% of the total charge.

composition was measured (see Table 7.1), before the addition of 1.86 or 18.6 mg/l UO2.(NO3)2.6H20 (equivalent to 1 and 10 mg of UO2+/I) which dissolved completely (samples were analysed for UO 2+ concentration). After standing, the pH of the synthetic groundwater was between 6.8 and 7.8, at which pH the aqueous carbonate species is HCO3.

2.4. Small column tests

Two small flow-through column tests were conducted on PANSIL to replicate the mode of contaminant exposure within a PRB. In these column tests 10 g of PANSIL was mixed with 20 g of acid-washed quartz sand to reduce the amount of PANSIL used for a particular column length. The PANSIL/sand mixture was placed in a 25-m1 glass column, which resulted in a pore volume of approximately 10 ml. The column was then saturated by the upward flow of distilled water. Once saturation was achieved the water was displaced with the upward flow of the desired contaminant solution at a constant flow rate of 1.6 ml/h (controlled by peristaltic pump). This flow rate gave a residence time for the contaminated solution within the column of about 6 h (a lower bound to typical groundwater residence times in a funnel and gate barrier). Effluent solutions were diverted through a spur at the top of the columns into a covered collection vessel. Effluent solutions were collected once a day and their volume and pH were measured prior to acidification for UO 2+ analysis. In the two tests:

(i) PANSIL was exposed to uranyl nitrate solution containing 30 mg/l UO 2+ buffered to pH 6 until the uranium concentration in the effluent equalled that in the influent. At that point, the column was leached with uncontaminated synthetic groundwater.

(ii) PANSIL was exposed to contaminated synthetic groundwater at pH 7 containing 30 mg/l UO~ + until the UO 2+ concentration in the effluent equalled that in the influent. At that point the column was leached with uncontaminated synthetic groundwater.

In addition a control test on acid washed sand only, using a 30 mg/l uranyl nitrate solution at pH --~ 5, was also run. In this test, initial breakthrough I occurred immediately, and influent and effluent UO 2+ concentration were equal when the average column loading was only 0.008 g/kg. Thus, it is assumed that the amount of uranium sorbed by the sand in the column tests can be ignored.

1 In this study initial breakthrough is operationally defined as a UO~ + concentration in the column effluent persistently above the analytical detection limit.


2.5. Column test with groundwater from P~cs in Hungary

A large column test (length 840 mm, cross-sectional area 100 mm 2) was conducted using PANSIL to treat the groundwater from a mine tailings disposal site near P6cs in Hungary (see Chapter 6 for further details). The flow direction in the column was from bottom to top, and the test was run at room temperature (20-25~ The test duration was 256 days, and the total volume of water passed through the column was 4288 ml at a flow rate of about 17 ml/day. The porosity of the PANSIL in the column was 40.7%, so the flow volume was equivalent to 126 pore volumes, with a residence time of 2 days. Thus, the superficial seepage velocity was about 0.4 m/day, which is at least an order of magnitude higher than the groundwater flow velocity at P6cs in Hungary.

The typical stable element chemical composition of groundwater extracted from near P6cs is given in Table 7.2. The pH of this groundwater is about 7. The uranium concentration at the site varies from several hundred lxg of UO~+/I to just over l mg of UO~+/I. The important ions in the groundwater used as the test column influent are shown in Table 7.3 (Na +, K + and CI- concentrations were not measured in the column influent). The uranium concentration in the column influent was 1.4 mg of UO~+/I.

The volume of effluent was monitored daily, and collected for analysis about once a week. The total uranium concentration in both the original groundwater and the column effluent were determined by a fluorimetric method, where the fluorescent intensity of a fused pellet containing the uranium from the solution sample is compared with that of a blank pellet. A URANUS fluorimeter manufactured by ALGADE (France) was used. Measurements are made on pellets containing 10% sodium fluoride and 90% sodium carbonate that had been fused at 910~ with the residue after evaporation of 0.1 ml of solution. The detection limit is about 10 txg/l. Usually triplicate measurements of the sample were made.

2.6. Geochemical modelling of the test solutions

To understand the variation in performance of PANSIL with pH and solution composition, numerical modelling has been undertaken using PHREEQE (version phrq96) equilibrium

Table 7.2. Typical groundwater composition at P6cs in Hungary (the uranyl concentration tends to vary spatially and temporally but is about l mg/l).

Na + K + Ca 2+ Mg 2+ C1- SO 2- HCO3

Concentration (mg/1) 70 17 168 53 45 337 484

Note: The error in charge balance is less that 0.5% of the total charge.

Table 7.3. Average measured composition of the groundwater used as the column influent (Na +, K + and CI- concentrations were not measured during the column experiment).

Ca 2+ Mg 2+ SO 2- HCO3 UO 2+

Concentration (mg/l) 163 67 385 595 1.4


geochemical modelling software and the CHEMVAL6 database. Three solutions were modelled, uranyl nitrate conditioned with l OO mg/l carbonate (as sodium carbonate) to replicate the effect of atmospheric CO2, the synthetic groundwater and the groundwater from P6cs. All these solutions were modelled under oxic conditions, the first with a uranyl concentration of 1O mg/l of UO~ +, the second with a uranyl concentration of 1 mg/l of UO 2+, and the Hungarian groundwater with a uranyl concentration of 0.8 mg/l of UO 2+. The equilibrium speciation of these solutions is shown in Figure 7.7 as a percent of the total UO~ + in particular species as a function of pH. As with the results of any geochemical model, the data presented in Figure 7.7 should be regarded as a guide to the likely speciation, as the precise values will depend on the geochemical database used.

In general, the uranyl speciation in the three solutions follows similar trends with pH, with the UO 2+ cation important below pH 5, the neutral UO2CO3 species important 5 < pH < 7, and anionic uranyl carbonate species dominant above pH 7. However, there are two significant differences between the solutions. Firstly, the neutral UO2SO4 species is important between pH 2 and 5 in both groundwater systems, and secondly, the neutral UO2(OH)2 species, present at circa-neutral pH, decreases in importance as the uranyl-to- carbonate ratio decreases. In the carbonated uranyl nitrate system, which has the highest uranyl-to-carbonate ratio of the three systems, the proportion of UO 2+ present as UO2(OH)2 exceeds 20% at pH 6 and persists at a percentage level to above pH 9. In the synthetic groundwater the proportion of UO 2+ present as UO2(OH)2 is just over 10% at pH 6 and is below 1% by pH 8. In the Hungarian groundwater, which has the lowest uranyl- to-carbonate ratio of the three systems, the proportion of UO~ + present as UO2(OH)2 peaks at less than 5% at pH 5 and is below 1% by pH 7.

2.7. Performance of PANSIL

The sorption isotherm for UO 2+ on PANSIL at pH --- 4.5 is shown in Figure 7.8(a) and (b), and that for UO22+ on PANSIL in the presence of an equal molar concentration of Pb 2+ (at about the same pH) is shown in Figure 7.8(c) and (d). Below a certain limit PANSIL removes UO~ + from solution to below the analytical detection limit, and can do so in the presence of an equal molar concentration of divalent metal ions. At pH --- 4.5 this level was between 300 and 500 mg of UO~+/kg of PANSIL (the higher capacity was determined from a mixed UO2+/pb 2+ isotherm, probably due to small differences in pH in the particular batch tests that define the capacities). Above this limit PANSIL sorbs UO~ + more weakly, establishing equilibrium between the sorbed and aqueous phase. At high solution concentrations the total sorption capacity at pH 4.5 is around 5 g/kg for the UO 2+ solution, but about 3 g/kg in the mixed UO22+/pb 2+ solution. This sorption behaviour, where at low loadings PANSIL preferentially sequesters uranyl in the presence of lead, whereas at high loadings sorption is non-specific for UO 2+, suggests that PANSIL operates by two different mechanisms as contaminant loadings differ. As a working hypothesis it is suggested that the uranyl-specific mechanism involves ring formation between uranyl ion and the amino and oxime ligands on PANSIL, whereas the non- specific sorption mechanism is probably electrostatic involving locations on the polymer or sand exhibiting a local negative charge.

The small column tests showed that when exposed to uranyl nitrate buffered to pH 6, initial breakthrough occurred at column loading of about 1.5 g of UO2+/kg of PANSIL,


(a) 100

90

80

70

6O

O D 50 o~

40

30

20

10

O

(b) too 90

80

70

60

O 50

40

30

20

10

O

(C) 100

90

8O

70

60

O 50

40

30

20

10

O

O 2 2 ~ + ~ UO2(CO3)~l L I " " ' . . . .

~ uo~/co~ ..',.. ,,

u o~c% / .,,/ oo~o.,~\ \ \ : :,.

uo~Io.)~ \ -"%'~ - ' ' uo o.

pH

~ll~ ~ § . u%(co3)4- �9 ~" "~', / t

.~ . -~- ' \ ;" ' , , UO2SO 4 /

! ~'P ./ ;- uo2(c%) ~ /" ~ , , .' " :' ., UO2CO3-" ' /

/ ~ N , . : : ', / U O 2 ( S O 4 ) 2 - ~ . -" ~,,'~ ~, "-..~UO2(OH ,2\ UO2'OH'3

.~1) i --~-' ~ ." ",. >, ..... -.._.__k

2 4 6 8 10

pH

Vo:. X , ?.so,

' / , ' ~ : '- uo co -", . , uo~(c%)~-

/ UOzOH + , , .. ' ~~ ( . \ : : f ', i %,, . I- . L ,* ~,

0 2 4 6 8 10

pH

Figure 7. 7. Geochemical modelling using PHREEQE and the CHEMVAL6 database. (a) Uranyl nitrate containing l O mg/l of UO~ + and 1OO mg/l of carbonate, (b) the synthetic groundwater containing l mg/l of UO 2+, and (c) the groundwater from P6cs in Hungary containing 0.8 mg/l of UO 2+.


(a) 6

5 t3~

4 v " o

.(3 3 " o < +,, 2 o E3

4 ' " �9

t" 200.00

0 0.00

�9 experimental data Langmuir model

| |

400.00 600.00 800.00 1000.00

(b) 1

0.8 o ) v

" o

0.6 O (/)

+< 0.4 O,l O~l o

0.2

/. /.

./ ./ /

0 o.oo 2doo 46.00

Conc. in Soln (mg/I) Conc. in Soln (mg/I)

(c) 6

4 O

O O " o

43 3

< o + 2 J o 0 / o O

1

0 o 2bo

o experimental data Langmuir model

0.8

)

0.8 (

0.4 ~/

0.2 ]

0 0

o/

t 4bo 6bo 8bo 1000 20 40

Conc. in Soln (mg/I) Conc. in Soln (mg/I)

Figure 7.8. Sorption isotherm for UO 2+ on PANSIL in (a) and (b) uranyl nitrate at pH 4.5, and (c) and (d) equal molarities of uranyl nitrate and lead nitrate at pH 4.5 (the initial sorption behaviour in both solutions is shown on expanded scales on the right).

and influent and effluent UO 2+ concentrations were equal when the average column loading was 2.8 g/kg (see Fig. 7.9). Subsequent leaching with uncontaminated synthetic groundwater at pH 7.5 initially liberated high concentrations of UO 2+ (probably due to divalent groundwater cations displacing the weakly sorbed U02+), but in the long term the PANSIL retained about I g of uoZ+/kg. When exposed to "synthetic groundwater", initial breakthrough occurred at an average PANSIL loading of about 1.1 g of UO~+/kg, and influent and effluent UO 2+ concentrations were equal when the average column loading was about 1.7 g/kg. When subsequent leaching with uncontaminated synthetic groundwater at pH 7.5 was undertaken, the effluent UO 2+ concentration showed no spike, instead it gradually declined, until in the long term the PANSIL retained about 1.5 g of uoZ+/kg (the decrease being small probably because there is significantly less weakly sorbed UO 2+ at high uranyl loadings when there is competition from divalent groundwater cations during initial exposure).


(a) 3-

2.5

0

o 1.5 o

0.5

-" C/Co ~ ~~176176 'k,

�9 " o UO 2 sorbed �9 -

Commenced leaching with r uncontaminated synthetic I I

" _ ~ # ' groundwater ~ ~ ~

J t Initia' ._.1

"Z.._._ 7" O 50 100 150 200

2.5

2 '~ O3 v " 0 (l)

1.5 0 CO

04 1 o

0 250

0.5

Volume passed (ml/g)

(b) 3.0

2.5

2.0 0

o 1.5 0

1.0

0.5

O.O

t - C/Co o UO 2 sorbed

Initial breakthrough

Commenced leaching with uncontaminated synthetic groundwater

3.0

2.5 ID')

2.0 v " 0 (l)

1.5 0

04

1.0 o E)

0.5

~ , T , O.O 0 20 40 60 80 100 120 140 160


Figure 7.9. Column tests on PANSIL using (a) uranyl nitrate solution at pH 6 and (b) synthetic groundwater at --~ pH 7.5 (both solutions contained 30 mg/l U02+).

Taken together, the sorption isotherm and small column test data indicate that the sequestration capacity of PANSIL increases with pH from 300-500 mg of UO~+/kg at pH 4.5, to about 1 g of UO~+/kg at pH 6 and possibly as high as 1.5 g of UO2+/kg at pH 7.5. It is likely that the PANSIL's increasing sequestration capacity with increasing pH is associated with the increasing ease with which oxime oxygen can undergo metal-assisted deprotonation, which increases the stability of the sorbed complex (see Section D.1.4).

The batch exposure tests conducted with uranyl nitrate solutions buffered to a range of pH values show that PANSIL is effective over a pH range 4-8 , with optimum performance in the pH range 4.5-7.5 (Fig. 7.10). In these tests the PANSIL was loaded to 300 mg of UO2+/kg of PANSIL, which is less than its estimated sequestration capacity and presumably was bound strongly to the PANSIL. In the two stage, distilled water, batch leaching tests on PANSIL that was loaded with UO~ + at pH 6, the UO 2+ concentration in


(a) 1

o) 0.8 E t -

O

"-= 0.6 0 if)

�9 ~ 0.4

o 0.2 E)

t t

Test Duration 24 hrs Liquid:solid ratio 30:1

E] . A ~ = 1 , == m ,

Final pH

(b) 10

~" 8 E -= 6

O

t - . m

4 ( I . ) n

o D 2

Test Duration 24 hrs Liquidsolid ratio 30:1

u

o

0 ;, 6 12 Final pH

h rip ~ Groundwater ! i

8 10

Figure 7.10. Variation of uranium sorption with pH by PANSIL in uranyl nitrate solutions containing (a) l mg/l, and (b) 10 mg/l of UO 2+. The mean (_ l standard deviation) of (a) 15 tests and (b) 7 tests with synthetic groundwater (C3) are also shown.

the leachate from the first stage was well below the detection method for the analytical method employed, and was undetectable in the second stage of these tests. These tests confirmed that, below its sequestration capacity, PANSIL strongly binds UO 2+.

The results of the batch tests using "synthetic groundwater" are also shown in Figure 7. lO. The maximum sequestration capacity of PANSIL for UO~ + at a neutral pH is about 6 mmol/kg (1.5 g of UO~+/kg). The synthetic groundwater initially contained 0.8 mmol/l (31.4 mg/l) of calcium, 1.8 mmol/l (43.0 mg/l) magnesium, and either 4 or 40 txmol/l (l or l O mg/l) of UO 2+, which at an L:S ratio of 30 is equivalent to 78 mmol of divalent cations per kg of PANSIL. Thus, the total concentration of divalent cations in the synthetic groundwater far exceeded the maximum UO 2+ sequestration capacity of PANSIL. Therefore, it is concluded that PANSIL can preferentially sequester UO 2+ from typical groundwater systems.

The longer duration pH controlled batch exposure tests showed that PANSIL reacts slowly in alkaline conditions, with amount sorbed decreasing with increasing pH, but after

3.5 Initial UO 2+ conc. 10 mg/I Liquid:solid ratio 30:1

3 (3t)

E v 2.5 u O 09 2 .=_.

_.e 1.5 ,,5+, 0 D 1

0.5

/

hrs ! hrs[ hrs /

1'0 12


pH

Figure 7.11. Effect of test duration on UO 2+ by PANSIL at high pH.

96 h it had sorbed significantly higher amounts of UO 2+ compared to 24 h even at p H l l (see Fig. 7.11). The leachate from the first distilled water leaching step on the PANSIL exposed for 96 h contained l mg of UO~+/1 when the pH was 8 during exposure, with the leachate concentration increasing slightly with pH to 1.3 mg of UO2+/I when the pH was 1 l. The leachate from the second leaching step contained 0.6 mg of U022+/1 when the pH was 8, increasing with pH to l mg of UO~+/I at pH 1 l. The UO 2+ concentration in leachate from the third leaching step was below the detection limit for all the experiments. Thus, at pH 8, ---82% of the UO 2+ in the original solution was strongly sorbed to the PANSIL, whereas at pH 11 only --- 48% was strongly sorbed to the PANSIL.

Two factors are thought to explain the pH dependence of UO 2+ sequestration by PANSIL in the batch exposure tests. At low pH, protonation of the lone pair of electrons on the nitrogens and oxygens effectively blocks the active sites on PANSIL thus retarding binding. At circumneutral pH it is the speciation of the uranyl ion that controls the effectiveness of the interaction, primarily by controlling the rate of strong reaction.

In the 24-h batch tests using uranyl nitrate solutions that were in equilibrium with atmospheric CO2, PA2qSIL was effective up to a pH of between 7.5 and 8. At these pH values the geochemical modelling indicates that the dominant uranyl species are either neutral or anionic (see Fig. 7.7). Likewise PANSIL was effective with the bicarbonate buffered synthetic groundwater system at pH--~ 7.5 where, again, the dominant uranyl species are either neutral or anionic. Thus, as the amino and oxime ligands on PANSIL are unlikely to interact effectively with anionic uranyl carbonate species due to repulsion of the lone pairs, it is deduced that PANSIL is sequestering the neutral uranyl species.

In the extended duration batch tests using uranyl nitrate solutions PANSIL was still reasonably effective in the alkaline range (pH 8- l l), but reaction rates were far slower than at pH 6. There seem to be two possible mechanisms by which PANSIL could operate in this range. Either PANSIL can sorb anionic species (probably electrostatically to locally


slightly positive sites on the PANSIL), or the amino and oxime ligands sequester neutral species (albeit present at very low concentrations at high pH) and thus, as equilibrium between the various uranyl species in solution is continually re-established, PANSIL can slowly (over a period of days) sequester significant amounts of uranium. The performance of PANSIL in alkaline conditions is compatible with a combination of these mechanisms operating. The extended time periods required to reach equilibrium (> 96 h; Fig. 7.1 l) are compatible with sequestration of minor species. In addition, the retention of significant amounts of uranium after three leaching steps suggests that much of the uranium is strongly sequestered, presumably by bidentate amidoxime ligands. However, the fact that a proportion of the UO 2+ initially sorbed to PANSIL was subsequently leached with distilled water indicates that some uranyl is weakly sorbed, which may represent weak physisorption of anionic uranyl carbonate species to the non-specific PANSIL surface.

In the large column test with groundwater from P~cs, where the influent uranyl concentration was approximately 1.4 mg of UO2+/I, breakthrough occurred when the column loading was about 35 mg of UO2+/kg of PANSIL (prior to breakthrough the effluent uranyl concentration was typically less than 0.05 mg of UO2+/I, and no individual measurement was greater than 0.13 mg of UO~+/I; Fig. 7.12(a)). The pH of the column effluent was about 8.1 throughout the test (Fig. 7.12(b)).

It may seem surprising that the capacity of PANSIL measured in column tests with the groundwater from Prcs is more than an order of magnitude lower than the capacity measured with synthetic groundwater, given the relatively small differences between the synthetic groundwater and the Hungarian groundwater (the former contained 30 mg of U022+/1 and 180 mg/l of carbonate at pH --- 7.5, the latter contained 1.4 mg of UO2+/I, ---600 mg/l of carbonate at pH 8.1). However, both these solutions are mildly alkaline, so PANSIL must operate under these conditions by sequestering neutral species. The geochemical modelling (Fig. 7.7) indicated that the lower uranyl and higher carbonate concentrations of the Hungarian groundwater probably resulted in a significantly lower UO2(OH)2 concentration in mildly alkaline conditions than in the synthetic groundwater. It is, therefore, suggested that the resultant change in speciation caused a far slower rate of reaction in the Hungarian groundwater. Thus, a residence time of 6 h was more than sufficient to remove most of the uranium from the synthetic groundwater (where about 1.5 % of the uranyl is calculated to be UO2(OH)2), but a residence time of 2 days was only just sufficient to successfully treat the Hungarian groundwater (where less than 0.05% of the uranyl is calculated to be UO2(OH)2). Breakthrough probably occurred in the Hungarian groundwater column because the sequestration capacity of the PANSIL near the inlet end of the column became exhausted, reducing the effective residence time of the column.

The extended residence time required to treat the Hungarian groundwater does not necessarily mean that PANSIL cannot be used to treat that solution because a PRB can, within reason, be designed to achieve the required residence time. The groundwater flow velocity at Prcs is more than an order of magnitude lower than used in the column test, but as PANSIL is likely to be used within a "funnel and gate" type barrier (where the groundwater is channelled through a reactive zone), the flow velocity used in the column test is of the correct magnitude for typical funnel and gate systems. However, the residence time within the barrier can be extended by increasing the barrier thickness; to a degree this can be achieved without increasing the amount of active material by adding inert filler, such as clean sand, to increase the volume of the reactive zone.


(a) 1.5

< 1 E

o D 0.5

Initial breakthrough

5 10 15 20


|

25 3O 35

(b) 9

"1- ,', 8

10 15 20


, |

25 30 35

(c) 8oo

6OO

E b~4OO

I 200

5 10 15 20 25 30


35

(d) 200

150

E 100

t ~

50

z~ A

I zx Infuent groundwater concentration �9 Column effluent concentration I

O 5 10 15 20 25 30 35


Figure 7.12. Column tests on PANSIL using the groundwater from P6cs in Hungary containing --- 1.4 mg/l UO 2+.


2.8. Durability of PANSIL

End-over-end agitation of PANSIL in various test solutions during batch testing (physically a more aggressive environment than would be encountered within a PRB) produced very little spalling of the modified PAN coating and no detectable damage to the sand support matrix. Thus, PANSIL will be sufficiently durable for any PRB application. Also, the use of sand as the support matrix means that it would be straightforward to tailor the permeability (hydraulic conductivity) of PANSIL to a particular PRB application by varying the sand size. During testing, the measured changes in solution pH were small, and no precipitation or other adverse chemical changes were detected. Indeed, as PANSIL works by selective sorption of the target species, it tends not to change the overall solution chemistry much, and thus precipitation of by-products such as carbonates that can clog PRBs should not occur.

3. Comparison tests with ion exchange resins

The performance of PANSIL was compared with that of two commercially available active materials by subjecting these materials to a similar programme of testing to that described above for PANSIL. The commercially available materials were Amberlite | IRN 77, a cation exchange resin with sulphonic acid functional groups, and Varion AP (manufactured in Hungary by Nitrokemia), an anion exchange resin with functionality based on -2,N-dimethyl pyridinium groups. The anion exchange resin has been used within the PEREBAR project with the groundwater from P6cs, Hungary.

3.1. Test programme for the ion exchange resins

Both resins were subjected to a series of batch tests similar to those performed on PANSIL (the sorption isotherms were determined using an L:S ratio of lOO: l). The cation exchange resin was supplied in an H + form and was tested both as supplied, and after conditioning with NaCI to convert it to a Na + form (to replicate the situation where the resin' s capacity to buffer the solution is exhausted, which occurs rapidly in groundwater). The anion exchange resin was conditioned with NaOH before testing (see Barton, 2003), and the uranyl nitrate test solution was conditioned with l OO mg/l of bicarbonate (sodium bicarbonate). Both resins were also subjected to small flow-through column tests to replicate the mode of contaminant exposure within a PRB. In these column tests 1 g of ion exchange resin was mixed with 29 g acid washed sand and placed in a 25-m1 glass column. The test procedure was the same as for the small PANSIL column tests. Once again the flow rate was 1.6 ml/h which gave a residence time of about 6 h. Four different test scenarios were investigated:

(i) The unconditioned cation exchange resin exposed to uranyl nitrate solution containing lOO mg/l UO 2+ buffered to pH 6. This test was abandoned after 12 weeks without UO 2+ being detected in the effluent. Throughout this time the effluent pH was 2.5.

(ii) The unconditioned cation exchange resin exposed to contaminated synthetic groundwater at pH 7 containing 30 mg/l of UO 2+.


(iii) The conditioned anion exchange resin exposed to uranyl nitrate solution conditioned with 100 mg/l of carbonate (sodium carbonate), containing 100 mg/l of UO 2+ and buffered to pH ~ 6.5 with HNO3. When the UO 2+ concentration in the effluent equalled that in the influent, the column was leached with uncontaminated synthetic groundwater.

(iv) The conditioned anion exchange resin exposed to synthetic groundwater at pH7 containing 30 mg/l of UO 2+.

3.2. Preliminary tests on the ion exchange resins

Increasing duration batch exposure tests where the L:S ratio was 30:1 and the solution initially contained 10 mg of UO2+/I indicated that an exposure time of about 2 h was sufficient for both the conditioned cation exchange resin to reach equilibrium with uranyl nitrate solution at pH--~ 5.5, and for the conditioned anion exchange resin to reach equilibrium with uranyl nitrate solution at pH --~ 7.2 (see Stewart et al., 2003, for details).

The sorption isotherm for the cation exchange resin indicates that when the solution concentration is 0.4 mg of U02+/1 at pH ~ 4, about 160 g of UO 2+ is sorbed per kg of resin (Fig. 7.13(a)). The isotherm for the anion exchange resin (Fig. 7.13(b)) indicates that when the solution concentration is 0.25 mg of UO2+/1 at pH ~ 5, about 20 g of UO 2+ is sorbed per kg of resin. These values may be taken as indicative of the sorption capacities of the resins in the absence of competing species. However, because ion exchange resins are not specific for uranium, these values are not necessarily a good guide to their likely sorption capacities in groundwater where competing species will be present.

3.3. Performance of the cation exchange resin

In the batch exposure tests on the unconditioned cation exchange resin, the resin achieved very high degrees of UO~ + removal regardless of initial solution pH, but did so by buffering the pH to below 4 thus allowing sequestration of the UO 2+ (aq) ion which is expected to be a dominant species at this pH. When the cation exchange resin was pre- conditioned with NaCI, it was very effective at pH values below 6.5, but was decreasingly effective above this pH and wholly ineffective at high pH (see Fig. 7.14). The conditioned cation exchange resin was wholly ineffective at removing uranium from the synthetic groundwater at pH ~ 7.

In the column test on the unconditioned cation exchange resin exposed to synthetic groundwater containing 30 mg/l UO 2+, initial breakthrough occurred when the amount of uranium sorbed was approximately 7.5 g/kg of resin (see Fig. 7.15). It should be noted that no datum point is shown in Figure 7.15 for the moment at which maximum UO~ 2+ sorption occurred (the point where influent and effluent concentrations are equal). Prior to initial breakthrough the effluent pH was 2.5, but after breakthrough it was 7.5. Thus, it is deduced that uranium breakthrough occurred because the resin's capacity to buffer the synthetic groundwater pH to a value where cationic uranium species predominated was exhausted. Shortly after breakthrough the effluent uranium concentration exceeded the influent concentration, which suggests that the unbuffered pH 7.5 synthetic groundwater caused the release of sorbed uranium presumably as neutral or anionic species. The results of this


(a) 600

~. 500 E}')

400 " o (D

..Q ~ 3OO

~ 200 �9 D 100

#

Ii,.

2'0 4'0 6'0 8'0 100 Conc. in solution (mg/I)

(b) 100

80

-o 60 .121

O

~ 40

O D 20

0 200 400 600 800 1000 Conc. in solution (mg/I)

Figure 7.13. Sorption isotherms for (a) the unconditioned cation exchange resin in uranyl nitrate solutions at ~pH 3.8, and (b) the conditioned anion exchange resin in uranyl nitrate solutions buffered with sodium carbonate at pH --~ 5.

test are a sharp contrast with those of the aborted column test using uranyl nitrate at pH 6, where a resin loading of 320 g/kg was achieved without breakthrough (the effluent pH was 2.5 throughout) presumably reflecting the presence of bicarbonate as a buffer in the synthetic groundwater.

Thus, the cation exchange resin has a very high uranium sorption capacity in acid conditions where the UO 2+ species dominates. However, in the natural environment this capacity will be significantly reduced by competition from other cationic species (UO 2+ can be below Ca 2+ and Mg 2+ in the cation replaceability series for some sorbents; Peters et al., 1974). In neutral groundwater systems, where neutral and anionic uranyl species predominate, it is ineffective. The unconditioned H + form of the resin can still remove a limited amount of UO 2+ from typical groundwater solutions, but does so by sorbing the natural groundwater cations and releasing H +, and thus reduces the solution pH (changing the physicochemical conditions in the aquifer after the barrier). This moves the uranyl equilibrium towards cationic UO2+(aq), which then sorbs to the resin. Once the sorbed H + from the unconditioned resin has been displaced the effluent pH can recover to a value

New barrier materials ' 175

10

~" 8 13}

t -

5 6 r

._=

4 m

+~ 0 S3

2

0

I Test Duration 24 hrs Liquid:solid ratio 30:1

0 4 6 8 10 12 Final pH

i

2

Figure 7.14. Variation of uranium sorption with pH by the conditioned cation exchange resin in uranyl nitrate solutions containing l O mg/l of UO 2+. The mean (_ l standard deviation) of four tests with synthetic groundwater is also shown (C]).

3 ; C/Co

~, UO 2 sorbed

30

25

O'}

20

0

o 1.5- (.)

1-

0.5

0 0

O') v

15 Initial o

cq

o ~ --"- 10 D

, ~ ~ 5

~ r 0

100 200 300 400 500 600 700 800 900 1000 Volume passed (ml/g)

Figure 7.15. Column tests on the cation exchange resin using synthetic groundwater at pH---7.5 containing 30 mg/l UO 2+ (note: no datum point was measured for the moment at which maximum UO 2+ sorption occurred).

close to the influent pH, where most of the important uranyl species are either neutral or anionic and do not interact with the cation exchange resin. In these conditions the cation exchange resin may gradually release sorbed UO 2+ due to competition by mass-action from natural ions in the groundwater (if the UO 2+ cation is below Ca 2+ and Mg 2+ in the cation replaceability series for IRN 77, then desorption could be rapid, as other large metal cations are readily replaced by other species with a stronger affinity for the resin; Rengaraj, 2001).


3.4. Performance of the anion exchange resin

The pH controlled batch tests showed that the anion exchange resin was effective at pH values above 4.5 (Fig. 7.16). Thus, the anion exchange resin was effective in neutral and alkaline conditions, where the modelling suggests there are significant concentrations of anionic uranyl species (see Fig. 7.7). It is hard to explain why the anion exchange resin should be effective below pH 5 where the modelling indicates that there are no anionic uranyl species (carbonate will out-gas as C02 in acidic conditions). However, it should be noted that data from the pH controlled batch tests are presented in terms of the final equilibrium pH, and the sample with a final pH of 4.5 had an initial pH of 5.2. The anion exchange resin was also very effective at treating U022 + contaminated synthetic groundwater (also shown in Fig. 7.16).

During the column test on the conditioned anion exchange resin using uranyl nitrate conditioned with sodium carbonate, the influent pH was in the range 6.5-6.9 and the effluent pH was typically half a unit higher than the influent. Initial breakthrough occurred when the resin loading was 80 g UO2+/kg, and the uranium concentration in the effluent was equal to that in the influent when the resin loading was 140 g UO~+/kg (see Fig. 7.17(a)). Subsequent leaching with uncontaminated synthetic groundwater at pH --- 7.5 resulted initially in a high UO~ + concentration in the effluent, and then a rapid decrease to a small fraction of the concentration used during contamination. By the end of testing the UO~ + concentration in the effluent was below the detection limit of the measurement method.

In comparison, initial breakthrough in the column test on the anion exchange resin using synthetic groundwater at pH---7.5 occurred when the resin loading was 50 g UO2+/kg, although the effluent concentration remained low (between 0.2 and 0.6 mg/l UO~ +) until the resin loading was 90 g UO~+/kg (see Fig. 7.17(b)). The uranium concentration in the effluent was equal to that in the influent when the resin loading was 120 g UO2+/kg.

10

8 m

E v

_= 6 O

._= 4

(l,} m

o :~ 2

Test Duration 24 hrs Liquid:solid ratio 30:1

�9 �9 �9 A & , A l l l i l

0 2 4 6 8 10 12

Final pH

, I - . . 1

Figure 7.16. Variation of uranium sorption with pH by the conditioned anion exchange resin in carbonated uranyl nitrate solution containing l 0 mg/l of UO 2+. The mean (+ l standard deviation) of three tests with synthetic groundwater is also shown (D).


(a) 3 �9 300

2.5

.

o

1.5. o

0.5.

0 0

,t C/Co

UO 2 sorbed Started leaching

~r

Initial ~ ~ ~ ~ breakthrough t , ~ " ~ " " - ~ :,,'-":, ~,",

- !

50

560 10'00 1s'oo 2000 . . . . . . , X " ~ ' ~

- 25O

-200 "O

150 ",.- O

100 o


(b) 1.5

1.251

1 O

o 075 o �9

0.5

0.25

�9 c/co

, ' , U O 2 sorbed . ~ , . , ~ ~ _ ~ ~ .,,. A

150

125

-100 "O (D

...Q 75 ~ 0

50 O

25

0 1000 2000 3000 4000 5000 6000 Volume passed (ml/g)

Figure 7.17. Column tests on the anion exchange resin using (a) carbonated uranyl nitrate at pH 6.5 containing 100 mg/l UO 2+ (after the influent and effluent UO 2+ concentrations were equal, the column was leached with uncontaminated synthetic groundwater), and (b) synthetic groundwater at pH--- 7.5 containing 30 mg/l UO 2+.

Thus, the anion exchange resin has a very high uranium sorption capacity (80 g/kg) in a pH neutral uranyl nitrate solution where neutral and anionic uranyl carbonate species (particularly U02(C03) 2-) dominate. However, this capacity is about a third lower in a groundwater system where there will be competition from other anionic species. Interestingly, the sulphate concentration in the synthetic groundwater was 2.6 mmol/l whereas the uranyl concentration was only l l O Ixmol/l. This suggests that the anion exchange resin has good specificity for U02(C03)22-. Nonetheless, as the uranium is sorbed to the surface of the resin by electrostatic interaction, it could be gradually released over time when the resin is leached by groundwater due to competition by mass action from the groundwater anions (in particular, divalent groundwater anions such as sulphate which is present in higher concentrations in the P~cs groundwater compared to our


synthetic groundwater). However, this characteristic does make it relatively straightforward to regenerate the resin for re-use.

4. Comparison of PANSIL performance with that of the ion exchange resins

In acidic solutions for which it is designed, the cation exchange resin is very effective at removing cationic UO~+(aq) from solution (although it is non-selective for that contaminant). However, in neutral groundwater systems where neutral and anionic uranyl species predominate, it is ineffective. In such solutions the unconditioned H + form of the resin can still remove a limited amount of UO 2+, but does so by causing an undesirable pH reduction. On the other hand, the anion exchange resin has a high capacity for anionic uranyl carbonate species and is, therefore, very effective in solutions more alkaline than pH 5. In this pH range it has good specificity for these uranyl species in the presence of aqueous sulphate and carbonate and is, therefore, very effective in neutral groundwater systems.

The new ligand-based active material, PANSIL, is effective at sequestering cationic and neutral uranyl species from solution when the pH > 4. However, the rate of sequestration decreases rapidly when the pH exceeds about 8, where neutral uranyl species are present only at very low concentrations. Thus, optimum performance is obtained in the pH range 4.5-7.5. PANSIL has very good specificity for uranyl species in the presence of typical groundwater ions, and once sequestered the UO 2+ is not readily leached from PANSIL. However, the uranium capacity of the pilot batch of PANSIL tested in this study is significantly lower than that of the commercially available anion exchange resin.

5. Other factors likely to effect the performance of PANSIL

Modified PAN has been shown to preferentially adsorb a range of heavy metals in the presence of alkali, alkali-earth and first row transition metals (Potts et al., 1990; Lin et al., 1993b; Rivas et al., 2000). It may even show greater affinity for some heavy metals (such as vanadium) than uranium. However, this is considered an advantage for its use in groundwater remediation, as it must be considered desirable to remove those heavy metals at the same time as the uranium, even if the amount of active material required for effective remediation is increased. Potentially more problematic is that modified PAN can sequester some iron species, which will affect its performance when treating groundwater with particularly high levels of iron in solution. However, high- dissolved iron concentrations in uranyl contaminated groundwater are only likely under acidic conditions, where PANSIL is, in any case, ineffective.

E. Technological applicability

The performance PANSIL and the ion exchange resins in contaminated synthetic groundwater is compared with the reported performance of other PRB materials in similar solutions in Table 7.4. The comparison is made in terms of the


Table 7.4. Comparison between the performance of the resin based materials with other PRB materials (after Bryant et al. (2003)).

Reactive material References pH Carbonate a Rd (mUg) (estimated) (mg/l)

Hydroxyapatite-based materials Pelletised bone charcoal Fuller et al. (2002b), ~ 7.2 400, 580 100-370

Joye et al. (2002) Crushed phosphate rock Fuller et al. (2002b) 7.3 580 30-125

Iron oxide pellets Joye et al. (2002) 7.2 400 25 PANSIL 7.6 180 > 1470 b Conditioned varian AP 7.4 180 > 1470 b Conditioned IRN 77 7.2 180 6

Data are presented from batch tests where the initial UO 2+ concentration was between 5 and 35 mg/l, the pH was between 7 and 8, and the L:S ratio was between l O:l and 100:1. a Final solution UO 2+ concentration was below the analytical detection limit. Therefore, a minimum Rd value was calculated from that detection limit (0.2 mg/l). b Between pH 7 and 8 the dominant carbonate species will be HCO3.

conditional distribution coefficient, Rd (mUg), which is the amount of a substance sorbed (Ixg/g) divided by its equilibrium concentration in solution (Ixg/ml). Generally, Rd is not a unique property of a particular sorbent/solution system, and usually varies with the test conditions (particularly the contaminant concentration). Thus, the test data used for comparison have been carefully selected to ensure that the contaminant concentration, pH and L:S ratio are similar to those used in this study. Also, PANSIL and the ion exchange resins remove UO 2+ by surface adsorption, and at the UO 2+ concentrations considered hydroxyapatite and iron oxide operate by similar mechanisms (Morrison and Spangler, 1992; Fuller et al., 2OO2a; Joye et al., 2002). Thus, the performance of all the materials reported in Table 7.4 depends on specific surface area, and so only similarly sized materials are compared.

Both the anion exchange resin and PANSIL compare very favourably with the other materials reported in Table 7.4, particularly when it is noted that the Rd values of the anion exchange resin and PANSIL are underestimated in Table 7.4 because, for calculation purposes, the final solution concentrations were assumed to be equal to the analytical detection limit (0.2 mg/l) rather than at a (more realistic) lower concentration. Indeed, the Rd values of the anion exchange resin and PANSIL are an order of magnitude higher than those reported for similarly sized bone charcoal and iron oxide pellets, and finer sized crushed phosphate rock. Unsurprisingly, for a material designed to adsorb cationic species from acidic solutions, the conditioned cation exchange resin was ineffective in the neutral synthetic groundwater.

Table 7.5 lists the approximate cost of various PRB active materials, including those used in this study. While the values in Table 7.5 should be taken only as a guide to the likely cost (the actual cost will vary with the cost of raw materials, exchange rates, the amount ordered, etc.), they do allow rough comparisons between the different materials. Despite the


Table 7.5. Typical cost of a range of reactive materials.

Reactive material Approximate c o s t a (US$/ton)

Elemental iron (Puls, 2002) Hydroxyapatite (Conca and Wright, 2003) Zeolites (Niederbacher et al., 2003) Activated carbon (Niederbacher et al., 2003) Bone charcoal (Bostick et al., 2000) PANSIL b Varion AP (Niederbacher et al., 2003) Amberlite IRN77 c

390 500 450

1280 2000

700 3300 4000

a Assuming IC ~ 1US$. b Estimated from cost of resin. c Cost of equivalent product.

limitations of these estimates, it is apparent that the ion exchange resins are about lO times more expensive than elemental iron and hydroxyapatite; about four times more expensive than PANSIL and activated carbon; and about twice as expensive as pelletised bone charcoal. However, the significantly higher Rd values of both the anion exchange resin and PANSIL appears to justify their higher cost (particularly as their Rd values were conservatively assessed). Further, the roughly 5O-fold higher capacity of the anion exchange resin, compared with PANSIL, measured in the flow-through column tests with synthetic groundwater at pH 7.5 would seem to make it the obvious choice for treating neutral uranium contaminated groundwater when carbonate is present.

Finally, it should be noted that the PANSIL tested during this project was from a prototype batch, where only a proportion of acrylonitrile groups was converted to amidoxime groups, and only partial coating of the sand with the polymer was achieved (see Fig. 7.1). It is believed that with better optimisation of coating process, a higher degree of sand particle coating can be achieved and that coating with a polymer that has a higher degree of functional group conversion may be possible, resulting in PANSIL with a significantly higher sequestration capacity. Thus, PANSIL may yet prove to be a cost-effective material for removing uranium from neutral carbonated groundwater.

F. Conclusions

Attempts have been made to synthesise new ligand-based active materials for groundwater treatment within a PRB following three distinctly different strategies. Of the three strategies, that based on coating an inert support particle with active polymer produced the most promising active material. That material, called PANSIL, is silica sand coated with a modified PAN resin (modified PAN is a polyacrylonitrile starting material where a proportion of the nitrile groups is converted to amidoximes).

PANSIL has been subjected to an extensive test programme to determine the conditions under which it works effectively, and to evaluate its performance in carbonated groundwater at intermediate pH. The following are the results.


�9 PANSIL should be sufficiently permeable for most PRB applications, and both the support matrix (quartz sand) and the modified PAN coating are durable.

�9 PANSIL is effective at sequestering cationic and neutral uranyl species when the solution pH is above 4, due to the stability of the polyacryloamidoxime-uranyl complex formed. However, the rate of sequestration decreases rapidly when the pH exceeds about 8 where the neutral uranyl species that complex with PANSIL are present only at vanishingly small concentrations. Under such conditions PANSIL is thought to sequester those neutral species, and can slowly accumulate significant amounts of uranyl, as equilibrium is continually re-established between the various uranyl species in solution. Thus, optimum performance is obtained in the pH range 4.5-7.5.

�9 PANSIL can preferentially sequester UO~ + from solutions that are typical of the groundwater from a mine tailings site. However, its performance is sensitive to the exact groundwater water chemistry, and small changes in the pH and the concentration of uranyl complexing ligands such as bicarbonate can dramatically affect the sequestration rate, which is an important design parameter for a PRB as it controls the barrier thickness required to treat a particular groundwater flow rate.

�9 The UO 2+ sequestration capacity of the batch of PANSIL tested was about 1.5 g of UO2+/kg at intermediate pH. Once sequestered by PANSIL the U022 + is not readily leached by distilled water or solutions typical of groundwater.

�9 PANSIL works by the selective sorption of the target species, and should not have much other effect on solution chemistry. It is, therefore, unlikely to cause the precipitation of by-products that can block porous barriers.

In neutral solutions with a composition typical of groundwater, the uranium capacity of the pilot batch of PANSIL tested in this study is significantly lower than that of the commercially available Varion AP anion exchange resin. However, PANSIL has excellent specificity for UO~ +, and its performance is unaffected by natural groundwater cations. Also, it should be possible to increase the uranium capacity of PANSIL by optimising the coating process.

Acknowledgements

The authors would like to acknowledge the help of Dr Nick Bryan, Radiochemistry Research Centre, Department of Chemistry, Manchester University, with the PHREEQE modelling, and the support and guidance of Dr T.P. Kee, School of Chemistry, University of Leeds.

References

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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 (4), 488-498.


Bryant, D.E., Stewart, D.I., Kee, T.P., Barton, C.S., 2003. Development of a functionalised polymer- coated silica for the removal of uranium from groundwater. Environ. Sci. Technol. 37, 4011-4016.

Conca, J.L., Wright, J., 2003. Apatite II to remediate soil or groundwater containing uranium or plutonium, 2003. Radiochemistry Conference, Carlsbad, NM, http://www.clu-in.org/conf/itrc/prb/pu = uapatite. pdf.

Fuller, C.C., Bargar, J.R., Davis, J.A., Piana, M.J., 2002a. Mechanisms of uranium interactions with hydroxyapatite: implications for groundwater remediation. Environ. Sci. Technol. 36, 158-165.

Fuller, C.C., Piana, M.J., Bargar, J.R., Davis, J.A., Kohler, M., 2002b. Evaluation of apatite materials for use in permeable reactive barriers for the remediation of uranium-contaminated groundwater, handbook of groundwater remediation using permeable reactive barriers. Academic Press, London, pp. 255-280.

Hall, P.O.J., Aller, R.C., 1992. Rapid small volume flow injection analysis for total CO2 and ammonium in marine and freshwaters. Limnol. Oceanography 37, 1113-1119.

Johnson, D.A., Florence, T.M., 1971. Spectrophotometric determination of uranium(Vl) with 2-(5-bromo- 2-pyridylazo)-5-diethylaminophenol. Anal. Chim. Acta 53, 73-79.

Joye, J.L., Naftz, D.L., Davis, J.A., Frethey, G.W., Rowland, R.C., 2002. Development and performance of an iron oxide/phosphate reactive barrier for the remediation of uranium contaminated groundwater, handbook of groundwater remediation using permeable reactive barriers. Academic Press, London, pp. 195-219.

Kabay, N., Katakai, A., Sugo, T., Egawa, H., 1993. Preparation of fibrous adsorbents containing anidoxime groups by radiation-induced grafting and application to uranium recovery from seawater. J. Appl. Polym. Sci. 49, 599-607.

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Lin, W.P., Lu, Y., Zeng, H.M., 1993b. Extraction of gold from Au(11I) ion containing solution by a reactive fiber. J. Appl. Polym. Sci. 49 (9), 1635-1638.

McCabe, D.J., Russell, A.A., Karthikeyan, S., Paine, R.T., Ryan, R.R., Smith, B., 1987. Synthesis and coordination chemistry of 2-(diethoxyphosphino)-pyridine and 2-(diphenylphosphino)pyridine N,P- dioxides - crystal and molecular structures of bis(nitrato)[2-(diethoxyphosphino)pyridine N,P- dioxide]dioxouranium(Vl) and bis(nitrato)[2-(diphenylphosphino)pyridine N,P-dioxide]dioxouranium(Vl)M. Inorg. Chem. 26, 1230-1235.

McComb, M.E., Gesser, H.D., 1997. Preparation of polyacryloamidoxime chelating cloth for the extraction of heavy metals from water. J. Appl. Polym. Sci. 65 (6), 1175-1192.

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 remediation. Environ. Sci. Technol. 26, 1922-1931.

Niederbacher, P., Nahold, M., Biermann, V., Simon, F.-G., Ludwig, S., 2003. Economic Feasibility of the Project Results. Report, PEREBAR Project, European Union (EVKI-CT-1999-00186).

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Puls, R.W., 2002. Permeable reactive barrier successfully treats plume at US coast guard facility in region 4, http://www.epa.gov/epaoswer/hazwaste/ca/success/uscg.htm.

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Chapter 8

Electrokinetic techniques

Gabi Gregolec, Karl Ernst Roehl and Kurt Czurda

A. Introduction

Permeable reactive barriers (PRBs) are subsurface constructions situated across the flow path of contaminant plumes. The geochemistry of reactive barrier systems is complex, especially in cases where very reactive materials are used which might lead to drastic changes of the geochemical conditions in the groundwater (e.g. elemental iron). Laboratory and field data have shown that the behaviour of major groundwater constituents such as, e.g. Ca, Mg, bicarbonate and sulphate, has a strong effect on the long-term effectiveness of the attenuation process within the barrier material since under certain conditions they tend to precipitate and form secondary minerals in the barrier matrix (O'Hannesin and Gillham, 1998; MacKenzie et al., 1999; McMahon et al., 1999; Vogan et al., 1999; B lowes et al., 2000; Klein and Schad, 2000; Phillips et al., 2000; Yabusaki et al., 2001).

Electrokinetic methods are increasingly being considered for remediation of contaminated clayey soils (Alshawabkeh et al., 1999; Haus et al., 1999; Haus and Czurda, 2000; Czurda et al., 200 l). In the field of geoenvironmental engineering, most of the work on electrokinetics was undertaken at the end of the 1980s and in the 1990s (e.g. Lageman 1993; Acar et al., 1992; Bruell et al., 1992; Probstein and Hicks, 1993; Probstein, 1994; Acar and Alshawabkeh, 1996; Alshawabkeh and Acar, 1996; Haus and Zorn, 1998; Haus and Czurda, 1999; Haus et al., 1999). For electrokinetic soil remediation, electrodes are placed in the ground and a direct current (de) electric field is applied, which induces movement of the contaminants to the electrode reservoirs. When an electric field is applied to a wet soil mass, different electrokinetic flows occur simultaneously. These electrokinetic flows include the flow of fluids, charged particles and ions towards the electrodes. The fundamental transport mechanisms caused by an electric field are (Fig. 8.1):

�9 electroosmosis (movement of liquid relative to a charged stationary surface) �9 electromigration (transport of charged ions or ion complexes in solution) �9 electrophoresis (movement of charged particles relative to a stationary fluid)

Electromigration and electroosmosis occur in fine-grained soils, whereas in coarse- grained soils electromigration and electrophoresis are dominant.

In addition to mass transport processes, chemical reactions take place at the electrodes. The principal electrode reaction observed is the electrolysis of water. At the cathode, water

184 G. Gregolec, K.E. Roehl, K. Czurda

Figure 8.1. Electrokinetic phenomena induced by an applied electric field (modified after Haus, 2002).

is reduced leading to the production of hydrogen gas and hydroxide ions. At the anode, water is oxidized and oxygen gas and hydrogen ions are generated. The electrokinetic transport of contaminants in soils may be accompanied by a number of other physicochemical processes, such as diffusion of dissolved species, adsorption/desorption of the species to soil particles, chemical reactions occurring in the bulk of the pore solution (e.g. mutual neutralization of H + produced at the anode and OH- produced at the cathode), buffeting of the induced pH change by the soil-fluid-contaminant system, precipitation of dissolved components (when their local concentrations exceed the solubility limits) and/or dissolution of precipitates.

The task of the present study was to evaluate the feasibility of electrokinetic methods to positively affect the long-term efficiency of PRBs. The general idea behind this study was the coupling of electrokinetic processes with PRB systems to reduce the amount of groundwater constituents, which might impair the barrier function by coating or clogging by precipitates, by electrokinetically fencing them off from the barrier inflow.

The feasibility of coupling electrokinetics with treatment zones within fine-grained soils has been documented in laboratory and field experiments. The so-called "Lasagna" process described by Ho et al. (1995, 1999a,b) consists of a layered configuration of electrodes and treatment zones. The contaminants are electrokinetically transported into the treatment zones. Laboratory and field experiments with the "Lasagna" technique showed promising results, for instance for the removal ofp-nitrophenol (PNP) from clayey soils using activated carbon in the treatment zone and trichloroethene (TCE) using activated carbon and elemental iron. Yang and Long (1999) found that the addition of a Fenton-like process with scrap iron powder was beneficial to the electrokinetic remediation of a sandy loam saturated with an aqueous solution of phenol.

Figure 8.2 shows the main possibilities for combining electrokinetic techniques with PRBs. Considering the long-term behaviour of reactive barriers, the application of an electrokinetic fence upstream of the PRB appears to be a promising and practicable approach. The general aim is to reduce the concentration of groundwater constituents that

Electrokinetic techniques 185

Figure 8.2. Possibilities for the combination of electrokinetic techniques and permeable reactive barriers.

might impair the barrier function from the groundwater entering the reactive barrier material.

B. Scope and approach

PRBs are used for the treatment of contaminated groundwater plumes and therefore are usually placed in aquifers which consist of coarse-grained sediments having significant effective porosity and hydraulical permeability. The question that arises when attempting to fence off groundwater constituents from entering the PRB is: can a charged species be sufficiently retarded by an electric field to stop it migrating with the hydraulic flow? Therefore the following aspects have to be considered:

�9 Electrokinetics in aquifer material. The literature shows that while many laboratory and field experiments have demonstrated the applicability of electrokinetics for in situ remediation of contaminated fine-grained soils (where electroosmosis may play an important role) there is not much information available on electrokinetic phenomena within coarse-grained soils (e.g. Kim and Lee, 1999).

�9 Combined electric and hydraulic gradients. No quantitative experience is reported concerning the electromigrative transport of charged species against a hydraulic gradient within aquifer materials or coarse-grained soils.

Thus, the fundamental electromigratory behaviour of ions under the combined influence of electric and hydraulic gradients in a coarse-grained sediment needed to be investigated. Therefore laboratory studies were conducted focusing on electromigration, which is defined as the transport of ions or ion complexes in solution. The work included the following tasks:

�9 set-up of laboratory test equipment at different scales (small scale, bench scale) for electrokinetic experiments with coarse-grained soil materials;


Table 8.1. Overview of the electrokinetic experiment series.

Set-up Experiment Material Hydraulic series gradient

Model Matrix contaminant solution

Electrokinetic l Sandy soil cell

Aquarium

2 Sandy soil 3 Sandy soil/

loess loam 4 Sandy soil 5 Sandy soil 6 Sandy soil

7 Sandy soil 8 Sandy soil 9 Sandy soil

l Sandy soil 2 Sandy soil/

Fe ~

NaC1 Distilled water

CaSO4 Distilled water Na2CO3 Distilled water

Anode ---, cathode NaCI Distilled water Anode ---, cathode CaSO4 Distilled water Anode ~ cathode Tap water Tap water only

only Anode ----, cathode Ca/MgCI2 Tap water Cathode ~ anode NaCI Distilled water Cathode---, anode Ca/MgCI2 Tap water

Cathode ~ anode NaCI Distilled water Cathode ~ anode Ca/MgCI2 Tap water

�9 determination of the fundamental transport mechanisms and geochemical processes; �9 evaluation of the effect of electrokinetic processes on the performance of PRBs; �9 determination of the technical requirements and feasibility of technical implemen-

tation; �9 general recommendations for the assessment of the suitability of electrokinetic

techniques within PRB systems.

Additionally the possibility of mobilising contaminants from clay lenses into coarser material was studied in laboratory experiments. The different experimental series are compiled in Table 8.1.

C. Experimental set-ups and methods

1. Electrokinetic cell

Small-scale experiments were conducted in an electrokinetic cell (Fig. 8.3) with the soil core having a length of 22 cm and a diameter of 10 cm. The feature of this construction was that there were two pairs of electrodes installed. The active electrodes were placed within the electrode chambers and the passive electrodes were placed at each end of the soil body (Fig. 8.4). The voltage at the passive electrodes was regulated by the active electrodes and held constant at l OO V/m during the experiments. The advantage of this system is that the combined voltage drop across the electrode/solution interface, electrode reservoir and filter plate could be measured.

The soil was installed into the cylinder in layers of approximately 2 cm. Each layer was compacted mechanically. After waterproof sealing, the probes were installed and the soil


Figure 8.3. Electrokinetic cell with sandy soil. Note the five probes distributed along the soil core. The polyethylene compartments at the front sides of the soil body contain the active electrodes, the electrode chambers and the passive potential electrodes.

Figure 8.4. Schematic graphic of an electrokinetic cell (modified after Steger et al., 2001).


core was slowly infiltrated with water to assure complete saturation. For experiments where only an electric gradient was applied, the in- and outlets were closed and only the gas produced at the electrodes was allowed to escape out of the system. Assuming diffusion as negligible, mass transport then takes place only due to electromigration. For experiments combining electric and hydraulic gradients a water flow was generated by pumping solution through the soil core.

During all experiments the voltage at the electrodes and potential probes were recorded continuously by a computerised data acquisition unit. The eluates of experiments with hydraulic gradient were sampled at least daily and analysed for pH (WTW pH 197), redox potential (WTW pH 197) and electric conductivity (WTW LF 197).

After termination of the experiments the soil cores were dissected into 10 slices. For each slice, the water content and soil pH were determined following standard procedures. To analyse the ionic concentrations of the salts used in the model solutions, batch extractions with deionised water were carried out on the soil slices. The eluates of the batch extractions and the column eluates collected during the experiments were analysed for anions by spectral photometry (WTW, MPM 3000) and for cations by flame atomic adsorption spectrometry (Perkin-Elmer 3030B).

The electrokinetic cell set-up has been used previously solely for the treatment of clayey - low-permeable - soils. First runs with sandy - highly permeable - soils showed that several changes were necessary to prevent unintentional hydraulic flow and to get additional process information. The modifications included:

�9 use of coarser filter plates and special filter membranes; �9 enlargement of overflows and gas outlets; �9 construction of an adjustable base frame which allows the exact horizontal placement of

the electrokinetic cell; and �9 the addition of five potential probes across the soil core which allow the monitoring of

changes in pore fluid conductivity with time and place.

Table 8.1 gives an overview of the different experimental series conducted in the small- scale electrokinetic cells. The materials (soils, reactive material) and solutions used in the experiments are described in the following sections.

2. Aquarium

To scale-up the experiments a special bench-scale apparatus, the so-called "aquarium", was constructed (Fig. 8.5). The dimensions of the "aquarium" were 30 cm width, 60 cm length and 30 cm height. This experimental set-up allowed varying of the electrode configurations as well as the insertion of reactive materials. To prevent gas trapping and minimise layer formation the soil was mixed and built up under water. The soil was separated from the reactive material and the electrode chambers, respectively, by using a permeable filter membrane, which was stable even at extreme pH conditions. The electric field applied was 100 V/m and this was held constant by a set of passive and active electrodes. To measure the voltage development across the soil body two rows of potential probes were installed. Sampling of the eluate was done twice a day. The parameters analysed in the eluates were pH, redox potential and electric conductivity.


Figure 8.5. Bench-scale electrokinetic test apparatus called "aquarium" with a length of 60 cm, width of 30 cm and height of 30 cm. Left: without soil before the experiment; right: during an experiment.

Experiments with and without Fe ~ as a reactive material were conducted in this set-up (Table 8.1). After termination of the experiments the soil body was cut into slices perpendicular and parallel to the electrical field. The analytical programme of various chemical and physical parameters and ion concentrations in the soil samples was the same as for the small-scale experiments.

3. Model soils and reactive material

To study the fundamental electromigratory processes it was necessary to use a simple model system to avoid additional geochemical reactions such as adsorption and ion exchange. Therefore, a pre-cleaned, iron-free quartz sand (Table 8.2) was chosen as model soil for all experiments (DORSILIT | 9 S from Dorfner Ltd, Hirschau, Germany). The assumed hydraulic conductivity of this uniform medium sand, of 7.7/8.4 x 10 -4 m]s (assessed by the methods of Beyer and Hazen, respectively), is typical for sandy aquifers. Soil pH of the sand was 5.5 measured in deionised water. A natural loess loam was used in experiments examining electrokinetic transport of contaminants from low-permeable soil to high-permeable soil (Table 8.3). The loess loam was characterised as a silt with a clay content (<0.002 mm) of 13%, a hydraulic conductivity of 2.94 x 10 -8 m/s and an electroosmotic conductivity of 3 - 9 x 10 - l~ mZ/V s. Cation exchange capacity of the loess loam was 13.05 mol(eq)/lO0 g. As reactive material, elemental iron (cast iron grit, supplied by Gotthart Maier, Germany) was chosen (see Chapter 3).

Table 8.2. Chemical composition of DORSILIT | 9 S quartz sand (data given by supplier).

Oxides Percentage (%) Oxides Percentage (%)

Si02 89.40 K20 4.97 A1203 4.68 Na20 0.29 Fe203 0.05 CaO 0.03 TiO2 0.20 MgO 0.02


Table 8.3. Mineralogical composition of loess loam.

Minerals Percentage (%) Minerals Percentage (%)

Quartz 40 Illite 15 Calcite 12.0 Chlorite 10 Dolomite 5.5 Smectite 5 Feldspars 5 Kaolinite 5 Organics 2.5

4. Model solutions

To verify the correct functioning of the experimental set-up and to examine fundamental electromigration behaviour in the model soils, it was decided that simple solutions should be used as model solutions in the laboratory experiments. A O.0l M NaCI (Merck, 1.06404) solution was chosen because of its stability and low attenuation in soil. A O.Ol M CaSO4 (Merck, 1.02161) solution was used as sulphate-rich solution because dissolved sulphate is a common groundwater constituent and can cause unfavourable precipitates in reactive iron barriers. More complex solutions were chosen to simulate more realistic conditions. Tap water (Table 8.4) and tap water enriched with 0.002 M CaCI2 (Merck, 2382) and 0.002 M MgCI2 (Merck, 1.05833) were used as representatives for multicomponent systems and to determine possible interactions between various ions. Regarding the major ionic constituents, the composition of the latter was close to the groundwater composition at the case study site in Prcs, Hungary (Chapter 9). For experiments examining the electroosmotic transport of ions through low-permeable soil into high-permeable soil (Table 8.1) a 0.02 M Na2CO3 (Merck, 1.06398) solution was used.

D. Theoretical model

Under certain conditions the movement of solutes in the pore solution can cease completely when applying an electric field (Hicks and Tondorf, 1994; Acar et al., 1995;

Table 8.4. Selected parameters of the tap water of Karlsruhe (Schnell, 2001).

Parameter Unit Average (n = 19)

pH - 7.28 Electric conductivity IxS/cm 626 Oxygen content mg/l 5.80 Sulphate mg/l 76 Chloride mg/l 25 Bicarbonate mg/l 316 Calcium mg/l 116 Sodium mg/l 13.9 Magnesium mg/l 13.0


Zorn et al., 2000). These situations have been recognized as a result of the electrolysis products. Therefore, during electroremediation the anolyte and catholyte solutions are used to maintain contaminant removal. Without anolyte and catholyte chemical control, quasi-steady-state conditions could be established. In the steady state, ions produced at each electrode meet within the soil at a reaction plane (Dzenitis, 1997). By using inert electrodes, the electrolysis of water leads to the production of H + at the anode and OH- at 'the cathode. This results in the development of a low pH front migrating from the anode to the cathode and a high pH front migrating from the cathode to the anode. Where these fronts meet they react to form water. For instance, in the case of electrokinetic separation of NaCI two binary solutes develop: HCI at the anode region and NaOH at the cathode region (Fig. 8.6).

Mass conservation for the steady-state distribution of a dilute species i in the absence of homogeneous reaction gives for any point x between anode and cathode (Dzenitis, 1997):

OC i OC i O2Ci Vc OX -- --ZiuiF--~x E -+- Di -0--~ (8.1)

where c is the molar concentration (mol/m3), Vc the bulk convection velocity (m/s), Zi the valence, ui the mobility (mol s/kg), F the Faraday constant (96,487 C/mol), E the electric field strength (V/m), and D the molecular diffusion coefficient (m2/s).

Figure 8.6. Schematic diagram of electroremediation and binary electrolyte regions for the case when NaCI is the initial electrolyte and H + and OH- are the primary electrode products (after Dzenitis, 1997).


When only two charged species (H + and CI- for the anode region and OH- and Na + for the cathode region) with constant diffusion coefficients are involved, the electric field could be eliminated in Equation (8.1) assuming electroneutrality. According to Dzenitis (1997) the concentration distribution for one binary region is

C~C, C)2C, Vc Ox = D, ~ (8.2) ~x 2

where c, is a reduced ion concentration formed from the cation molar concentration c and the stoichiometric coefficients of dissociation u:

C , + ~ C ,

c, -- - (8.3) / ~ , + / J , _

D, is the binary diffusion coefficient (Probstein, 1994):

Z,+ u ,+D,_ - Z , - u , _ D , + D, = (8.4)

Z,+/~',+ -- Z , - ~ ' , -

The analytic solution of Equation (8.1) is given by Dzenitis (1997). Computed steady-state conditions for different convective velocities for NaCI separation are given in Figure 8.7.

To evaluate the experimental results, a numerical model was applied which couples the different mass transport mechanisms (electromigration, hydraulic flow and diffusion) with the electric field distribution (conservation of charge) and chemical reactions. The chemical reactions considered are the electrolysis of water at the electrodes and the dissociation and formation of water.

oek : , , o

-~ 5 ................ ~. ........................................ t - O ...~ ~ 4 (-.

o c 3 O O

.__ 2

E L _

o ] E

O O.O

I ' I ' I ' ' l l

......... Vc=-2"OE-O7 ............................................................... i ............ ~ . . . . . . . Vc=2.OE-O7 ~ , / 9

/

..................... - ..................................... V c - O ............................................................. U - - - I ~ ................... c o Y 4

............ .....2........2..:~ ................................................................................................................................

', ,o, ,~176 "" ~ i

0.2 0.4 0.6 0.8 4 .O

normal ized position from anode x' [m/m]

Figure 8.7. Steady-state concentration distribution for different convective velocities (negligible convective displacement).


E. Results

1. Small-scale experiments (electrokinetic cell)

The objective of these investigations was to understand the fundamental electromigration behaviour of ions under the combined influence of electric and hydraulic gradients in a coarse-grained sediment.

1.1. Electric gradient only

Initial experiments were conducted in a closed system without applying a hydraulic gradient. Thus changes in the ion concentration within the model soil (quartz sand) are solely due to electromigration and diffusion processes. The model contaminants were NaCI and CaSO4, respectively (Table 8.1).

According to their charge, Na +, Ca 2+ and H + are migrating towards the cathode while CI-, SO 2- and OH- are migrating towards the anode. Theoretical analysis and numerical simulations by Dzenitis (1997) postulate that in such a simple system steady-state conditions are reached where the current density is low and contaminant transport by migration is very slow. The electrolysis products H + and OH- meet within the soil at a reaction plane where they form water. This region is characterised by a low electric conductivity and high electric field strength. As a consequence the electrolyte of the system is separated by this reaction plane into two binary zones: the anode zone dominated by HCI]HzSO 4 and the cathode zone dominated by NaOH/Ca(OH)2.

Model contaminant NaCl Based on the development of current and voltage profiles it can be concluded that within approximately 50 h steady-state conditions were reached. The current showed an initial rise, which represents an increased electric conductivity within the soil resulting from the production of additional ions through the electrolysis of water. The subsequent steep decrease in current arises from the formation of water within the soil core, which leads to a zone of low electric conductivity. The formation of water is also documented by the development of the potential gradient within the soil core, as documented by the distribution of the potential traces of the five potential probes (Fig. 8.8). The voltage at probes 1-3 increased with time whereas voltage of probes 4 and 5 decrease with time. The difference in voltage between probes 3 and 4 show that in this region the potential gradient is highest. The electromigration of ions led to local changes of ion concentrations, which in turn resulted in local changes of the electric field. As the electromigration velocity of a species is proportional to the electric field, velocity is high where steep voltage gradients exist and vice versa.

These findings imply that a water front formed around a distance of 0.6 from the anode. The distribution of the concentration of the Na + and CI- ions shows that there was a zone depleted in ions whereas towards the electrodes the ion concentration gradient was very steep (Fig. 8.9). The experimental results coincided very well with the predictions achieved when using Dzenitis's (1997) model (Fig. 8.9).


Figure 8.8. Change of potential at the probes (normalized to the voltage applied between the cathode and anode) with time and place for an experiment with only the electric gradient applied. Note potentials at 0.0 and 1.0 distances to anode are measured at the passive electrodes (Gregolec et al., 2001).

Figure 8.9. Experimental and modelled distribution of Na + and C1- in soil under steady state conditions; concentrations normalised to initial concentrations of the compounds (Gregolec et al., 2001).


Model contaminant CaSO4 To verify the findings of the NaC1 experiments, the experiments were repeated under similar conditions but now using CaSO4 as a more realistic model groundwater. The results regarding the development of current, potential gradients and change in ion concentration coincided fully with the NaCI experiment. Ca 2+ and SO 2- were transported towards the opposite-charged electrodes. As a consequence two binary zones enriched in ions (H + and SO 2- at the anode and Ca 2+ and OH- at the cathode) separated by a zone depleted in ions formed, and steady-state conditions were achieved. The zone depleted in ions formed approximately mid-way between the anode and the cathode. This zone was characterised by a steep voltage gradient. This phenomenon was again predicted very well by the theoretical simulations based on the model of Dzenitis (1997), which thus validated the experimental results.

Model contaminant Na2C03 Clay lenses are a common problem with regard to the decontamination of aquifers as they can contain substantial amounts of contaminants which are, due to the intrinsic properties of clays like low hydraulic permeability, large specific surfaces and negative surface charge, only released very slowly and therefore cannot be treated efficiently by conventional technologies. On the other hand, electrokinetic processes, particularly electroosmosis, allow the transport of contaminants within a defined zone especially in fine-grained soils. Thus, this technique can be used for the transport of contaminants from the clay lenses into the aquifer where they can flow with the groundwater into the reactive barrier.

To investigate whether contaminant cations can be transported out of a low-permeable soil into a neighbouring sand layer, small-scale experiments were conducted. The electrokinetic cell was half filled with loess loam and half with sand and was saturated with distilled water. After saturation was achieved an electric field of 100 V/m was applied with the anode being applied at the side of the loess loam and the cathode at the side of the sand. During the experiment the anode chamber was rinsed continuously with 0.02 M sodium carbonate solution stored in a reservoir at the anode side.

At the cathode, sodium breakthrough was achieved within a few hours. The initially lower sodium discharge rate is probably caused by retardation of sodium due to sorption/desorption mechanisms within the clayey loam. After 25 h discharge rates were constant.

The total amount of effluent at the cathode increased almost linearly with experiment time and proves that an electroosmotically driven water transport from the loam into the sand indeed takes place. Thus sodium is transported towards the cathode not only by electromigration but also by electroosmosis. With regard to the remediation success this means that not only charged but also uncharged contaminants will be transported by electrokinetic processes.

In summary these results show that the theoretically predicted transport of water and species through low-permeable soil into high-permeable soil indeed takes place. Therefore electrokinetics can be utilised for the decontamination of clay lenses within an aquifer and provide enhanced treatment rates.


1.2. Combined electric and hydraulic gradient

To investigate the fundamentals of electrokinetic retention of ions against a hydraulic flow, experiments combining electric and hydraulic gradients were conducted. For all experiments the hydraulic gradient was adjusted to approximately O.OOI, which is typical for aquifers and was generated by pumping a solution through the soil core. During all experiments the voltage was held constant at 20 V, which gave a field strength of 1OO V/m. The model solutions used were NaCI and CaSO4 in distilled water, tap water and Ca/MgCI2 in tap water (Table 8.1).

For experiments with NaCI, CaSO4 and tap water as the model solution the hydraulic gradient was applied from the anode towards the cathode to examine the electrokinetic retention of anions. For the case of Ca/MgCI2 as well as NaCI the hydraulic gradient was applied from the cathode towards the anode to investigate effects of electrokinetic retention of cations (Table 8.1).

To facilitate mass balancing and to understand the basics of electrokinetic retention the soil core was first flushed with distilled water. To simulate the situation where there is a constant supply of ions, the experiment with Ca/MgCI2 was also carried out by flushing the soil core with that solution.

High sulphate concentrations are often typical for contaminated groundwater. Thus, the electrokinetic retention of sulphate was the objective of additional experiments, where CaSO4 was used as model solution.

Electrokinetic retention of anions First, experiments combining electric and hydraulic gradients were carried out with sand as model soil and NaCI as model solution. Before the experiment, the soil core was saturated and flushed with this solution.

Development of the voltage profile along the soil core and the pattern of the current with time differed significantly from the experiments without hydraulic gradient. The current started to rise significantly 16 h after the start of the experiment. The reason is the water front, which is transported towards the cathode chamber by the hydraulic flow. This is shown by the development of the electric gradient through the soil (Fig. 8.10). Once the water front reached the cathode chamber the resistance of the catholyte increased strongly. Consequently the applied voltage across the experiment was increased by the computerised experiment control system to maintain a constant potential gradient across the soil. By reducing the applied voltage to l0 V (from the initial 20 V), current and potentials within the soil core remained constant. Increasing the voltage towards the end of the experiment led to their rise again. Considering the potentials measured at the probes, gradients remained constant at lower applied voltage (compare O and 138 h potential traces in Fig. 8.10) whereas higher voltage causes local changes again.

This fact is also reflected by the development of the concentrations of Na + and CI- in the cathode solution (Fig. 8.1 l). When the electric gradients were high the electromigration velocity of CI- was high enough to counteract hydraulic gradients resulting in a decrease of concentration in the cathode solution whereas the Na + concentration increased by the combined effect of electromigration and hydraulic transport. These results indicate that the separation of ions also takes place when a hydraulic gradient is applied. When on the other hand the average hydraulic velocity exceeded electromigration,


Figure 8.10. Change of potential with time and place for experiment with hydraulic gradient (model solution: NaCI). Note potentials at 0.0 and 1.0 distances to anode are measured at the passive electrodes (Gregolec et al., 2001).

Figure 8.11. Development of the concentrations of Na + and CI- in the cathode solution with time; concentrations normalised to initial concentrations of the compounds (Gregolec et al., 2001).


the concentrations of Na + and C1- in the cathode eluate remained at a constant level, indicating that the input of ions was equal to the output.

It is inferred that during these experiments electromigration dominated at higher voltage gradients and hydraulic flow overcame electromigration when lower voltage gradients were applied. The results prove that it is possible to hinder charged species from moving with the groundwater flow by applying an electric field.

Further experiments were conducted with CaSO4 model solutions. To avoid the passivation of the active cathode in these experiments the soil core was saturated with CaSO4 solution but flushed with distilled water. The development of potential gradients across the soil core at different times of the experiment showed the same features as for NaCI system indicating again that a water front was formed and transported with the hydraulic flow towards the cathode. Once the water front reached the cathode chamber, linear voltage gradients across the soil body are again achieved, indicating an even distribution of SO 2- in the soil. The development of ion separation was reflected by the differing discharge of Ca 2+ and SO 2- at the cathode. Ca 2+ was transported by electromigration and hydraulic flow, both directed towards the cathode, and is almost completely discharged at the cathode (90% of the initial soil content). In contrast, electromigration of SO 2- is directed opposite to the hydraulic flow and thus the removal rate of SO 2- at the cathode is much lower (around 35% of the initial soil content). The discharge of both ions was high in the initial phase of the experiment but almost ceases after the water front has moved out of the soil body. Consequently, from that time onwards the electromigration of sulphate and the hydraulic flow are in balance proving that the retardation of sulphate against a hydraulic flow is possible.

To investigate if these results are also valid for multi-component systems, tap water was used as a model solution and the experiments were repeated under same conditions. The potential gradient across the soil body showed the same features as for the more simpler NaCI or CaSO4 systems. Considering the distribution of CI- and Na + across the soil body after the experiment it could be seen again that the majority of the major anions (such as CI, $04) remained in the soil while the primary cations (such as Na, Ca, Mg) have been leached almost completely from the soil core (Fig. 8.12). The electrokinetic retention of anions is obviously effective in more complex systems also.

Electrokinetic retention of cations The electrokinetic retention of cations was investigated by applying the hydraulic gradient from cathode to anode. The model solutions used were NaCI dissolved in distilled water and CaCI2 and MgCI2 dissolved in tap water. With the latter solution two types of experiments were carried out: one with flushing the soil core with distilled water, the other flushing the soil core with the solution itself.

The total discharge of ions from the experiment with NaCI showed exactly the same features as seen in the previously described experiments targeted at the electrokinetic retardation of anions. CI- was transported towards the anode by both electromigration and hydraulic flow and leached completely from the soil core (100% of the initial content in the soil core), whereas Na + transport by the hydraulic flow was completely prevented by an opposing electromigratory flux. Thus the discharge rate as well as the total amount of Na + discharged at the anode (37%) was much lower compared to CI-. Again, separation of ions took place and at the end of the tests electromigration of Na + was in balance with


Figure 8.12. Distribution of Na + and CI- as representative ions across the soil body after electrokinetic experiment coupled with hydraulic gradient from anode to cathode (concentrations normalised to initial concentrations of the compounds).

the hydraulic flow. Consequently no CI- but approximately 75% of Na + remained in the soil core.

To determine if electrokinetic retention of cations also takes place in the presence of competing ions, tap water enriched with Ca/MgCI2 was used as a model solution. The total discharge of Mg 2+ and Ca 2+ at the anode reached maximum values of 23 and 32%, respectively, of the initial soil core content.

The distribution of Ca 2+ and Mg 2+ across the soil core after the experiment showed that the major amount of Ca 2+ remained within the soil whereas only around 20% of Mg 2+ could be recovered. Furthermore, a great lack in the mass balances of these two cations was observed (12% for Ca 2+ and 60% for Mg2+). This phenomenon can be explained by the electrochemical reactions that occurred at the electrodes. The electrolysis of water at the cathode led to the formation of hydroxide, which was flushed with the hydraulic flow through the soil core towards the anode. Thus, an alkaline pH of 9-9 .5 developed in the soil during the experiment (initial pH in the soil: 7-7.5) resulting in the precipitation of Ca 2+ and Mg 2+ as hydroxides. The better mass balance for Na + in the previous experiments reflects the lower tendency of Na + to precipitate.

Thus it appears that a combination of electrokinetics with a reactive barrier could prevent the advection of substantial quantifies of unfavourable cations into the PRB, thus avoiding their precipitation within the barrier.

To verify these observations further experiments with Ca/MgCl2-enriched tap water were conducted. In these experiments the soil saturated with the Ca/MgCI2 solution was not flushed with distilled water but with the same solution to simulate constant ion supply. The linear increase in the discharge of chloride at the anode (Fig. 8.13) indicated that the volumetric flow rate during the experiment was constant over time and that chloride was


Figure 8.13. Total discharge of CI-, Ca 2+ and Mg 2+ at the anode with time in an experiment simulating constant ion supply.

not retained in the soil. The total discharge of the cations stagnated after 3 -4 days at 8-10% of the input concentration.

The post-test distribution of the cations across the soil core determined in batch tests with deionised water showed a recovery of Ca 2+ and Mg 2+ of only 19 and 12%, respectively. The high soil pH of 8- lO developed during the experiment suggested the formation of Ca 2+ and Mg 2+ hydroxides which could not be dissolved in the water batch tests. Nevertheless, additional batch tests with acidic solutions to recover potentially precipitated hydroxides yielded only a low recovery of 23% for Ca 2+ and 20% for Mg 2+. Inspection of the cathode after termination of the experiment showed that substantial amounts of precipitation, presumably of Ca and Mg hydroxides, took place around it (Fig. 8.14).

Regarding the application of the electrokinetic technique upstream of a reactive barrier, this indicates that a considerable amount of unfavourable cations can be hindered from entering the barrier. Thus, clogging and coating within the barrier can be reduced significantly. As the major amount of hydroxides precipitate directly at the cathode, precipitation within the soil is low. Thus, only small porosity losses and subsequent changes in hydraulic parameters can be expected.

2. Bench-scale experiments

Small-scale experiments with the electrokinetic cell showed very promising results regarding electrokinetic retention of both anions and cations against a hydraulic gradient. As a next step, experiments at bench-scale ("aquarium tests") were carried out (Table 8.1). To evaluate the applicability of this experimental set-up the first model solution used was


Figure 8.14. Cathode after the electrokinetic experiment coupled with hydraulic gradient from anode to cathode and simulating constant ion supply (Ca/Mg), showing precipitates of Ca and Mg hydroxides.

NaCI as its geochemical reactivity is low, and results can be compared with those achieved in the small-scale electrokinetic cell. By applying a hydraulic gradient from the anode towards the cathode the electrokinetic retardation of CI- was investigated.

Like in the small-scale experiments, steady-state CI- discharge at the cathode was achieved 2 - 3 h after the start of the experiments, indicating the balance of electromigration and hydraulic transport of CI-. The discharge of CI- at the cathode was 7-7.5% of the total amount in the soil. These observations were also supported by the distribution of CI- across the soil after the experiment. The majority of CI- remained in soil which proved the electrokinetic retention of CI-. Since the results of this bench-scale experiment coincided well with the small-scale experiments it could be concluded that a transfer of principle mechanisms to larger dimensions was achieved.

Further experiments were conducted with tap water enriched with Ca/MgCI2 as model solution. In addition, elemental iron was incorporated into the model as reactive material. To simulate realistic conditions, the electrodes were placed as an electrokinetic fence upstream of the reactive material (Fig. 8.15). The hydraulic flow was directed from the cathode towards the anode, to examine electrokinetic retardation of cations. The solution then passed the reactive material before it was collected in the effluent reservoir (flow was from left to fight in Fig. 8.15). The electric field acts only between the electrodes and thus only affects ions that are placed between the electrodes. Thus ions outside of the electric field are only transported by hydraulic flow and diffusion.


Figure 8.15. Bench-scale experiment with electrokinetic fence upstream of reactive material.

The discharges of Ca 2+, Mg 2+ and CI- into the effluent reservoir showed the same features as in the experiments conducted before including the small-scale experiments, indicating the separation of ions and the balance between electromigration and hydraulic flow. The relatively high total discharge of the cations (Ca 2+ 70%, Mg 2+ 55%) in the effluent is related to the fact that a distinct amount of these cations was present in the soil outside of the electric field and is transported by hydraulic flow and diffusion.

After the experiment the soil was eluted with both distilled water and an acidic eluate to recover potential precipitates. Note that the soil core was separated into three different zones, which were sand within the electric field, sand outside the electric field and reactive material outside of the electric field. These zones were reflected by the distribution of the cations across the soil after the experiment (Fig. 8.16).

Within the electric field the cations were retarded to almost 100%. The acidic batch showed that precipitation, especially of Mg 2+ species, took place, mainly in the vicinity of the cathode. Precipitation also took place within the elemental iron, but in the sand zone outside of the electric field almost no cations were retarded.

The distribution of soil pH after the experiment (Fig. 8.17) gave a further explanation for the distribution characteristic of the cations after the experiment. Between the electrodes pH was approximately 9.5, within the elemental iron close to 9, whereas in the sand in-between soil pH was around 6. Thus, precipitation tendency within the different zones was a result of pH development within the soil. The soil pH itself was a result of electrochemical reactions at the electrodes. At the cathode, hydroxide ions were produced


Figure 8.16. Distribution of Ca 2+ and Mg 2+ across the soil body after bench-scale experiment coupled with hydraulic gradient from cathode to anode (concentrations normalised to initial concentrations of the compounds).

which were flushed with the hydraulic gradient towards the anode. At the anode, hydronium ions were generated which neutralised the alkaline pH. In the elemental iron a more alkaline pH was generated due to redox reactions that typically take place within an elemental iron matrix (see Chapters 1-6).

F. Discussion and conclusions

l. Summary of the experimental results

The approach of hindering unfavourable groundwater constituents from moving with the groundwater flow into a reactive barrier by applying an electrokinetic fence upstream of the barrier has been studied. Small-scale and bench-scale laboratory experiments were conducted to understand the principles of electrokinetic retardation of charged species in highly permeable soils. The experiments showed that it is indeed possible to use electromigratory processes in such soils to counteract the hydraulic flow.

When an electric field is applied, ions present in the pore water are transported towards the oppositely charged electrodes by electromigration. As a consequence, two zones enriched in ions form near the electrodes, separated by a zone depleted in ions. In experiments conducted only with an electric gradient (i.e. without hydraulic flow), steady- state conditions developed as predicted by model calculations. In this case the electric


Figure 8.17. Soil pH before and after the bench-scale experiment coupled with hydraulic gradient from cathode towards anode (Ca/Mg).

gradient is small over a substantial region of the soil. Electromigration is slow and the electrokinetic retardation potential of ions is low. This phenomenon has to be considered if groundwater is poor in constituents and/or groundwater velocity is very slow.

Laboratory experiments combining electric and hydraulic gradients proved that it is possible to hinder the bulk of charged species from moving with the groundwater flow. Uncharged species will of course still be transported in the direction of the hydraulic flow. In the experimental soil/water model systems, a field strength of l OO V/m was required to hinder the transport of monovalent and divalent ions when the hydraulic gradient was O.OOl. Thus it appears to be possible to electrokinetically trap groundwater constituents that, if transported through the reactive material of a PRB, might impair the function of the barrier.

In particular the electrokinetic retardation of cations appeared to be very promising. Retardation of cations not only takes place by electromigration but also by precipitation due to hydroxide formation at the cathode. Batch experiments showed that the majority of precipitation takes place directly at the cathode. Thus, precipitation rates within the soil itself are low and consequently small porosity losses and changes in hydraulic parameters can be expected.

Bench-scale experiments reproduced the results of the small-scale experiments. Therefore it is very likely that the observed electrokinetic retardation of ions can be transferred into field scale.

Regarding the application of an electrokinetic barrier in aquifers it is essential to know the hydraulic conditions of the subsurface precisely in order to balance the movement of


ionic groundwater constituents and the pH changes induced by the electrodes. The results show that equilibrium of the various effects can be directly controlled by the applied electric voltage.

2. Practical aspects

2.1. Technical equipment and installation engineering

In the following section an overview of the technical equipment needed and practical aspects for the installation of an electrokinetic system are given. In principle at least one pair of electrodes and a DC power supply unit is needed to apply an electrical field into the ground.

Electrodes The basic requirement for the electrodes is electric conductivity. Generally two different types of electrodes can be used: plate electrodes and electrodes placed within a well. The former are used mostly in laboratory and the latter in field tests. In principle, the construction of electrode wells differs little from groundwater monitoring wells. To avoid dissolution of the anode as well as the formation of undesirable corrosion products the use of chemically inert materials, like graphite, coated titanium or platinum, is recommended (Haus et al., 2002). For the cathode, any electrically conductive, base-resistant material can be used, such as most ordinary metals, e.g. iron or steel.

Electro-technical data The space between the electrodes should not exceed 10 m to ensure a sufficient electric gradient (Zorn et al., 2000). The voltage gradient should be between 10 and 100 V/m to achieve sufficient electrokinetic transport velocities, but should not exceed 500 V/m for safety reasons.

In an electric field a current density between 1.5 and 3 A/m 3 is needed to generate effective electroosmotic transport processes. The required current density for electromigration is less as the electric force directly affects the charged species. In turn the voltage depends on the distance between the electrodes and the specific resistance of the soil, which can vary along several magnitudes. In an electrokinetic fence, the required voltage will principally be dependent not only on the groundwater flow velocity but also on the targeted groundwater constituents.

Process water In the electrode reservoirs the transported species can be enriched or extracted and decontaminated in conventional waste water treatment facilities. At the cathode, precipitation of calcium, magnesium and metals takes place mainly by the formation of hydroxides and/or carbonates. The precipitation takes place not only directly at the cathode but also in the cathode reservoirs. Before going into the field the electrokinetic behaviour of both main groundwater constituents as well as contaminants should be examined in laboratory tests. It is strongly recommended that these lab tests are conducted under model field conditions, including the original soil and groundwater composition as well replicating the correct hydraulic conditions.


2.2. Costs

Like all remediation technologies, and especially the innovative ones, the costs of electrokinetically assisted PRB are dependent on a multiplicity of parameters. The cost of electrokinetic systems given in the literature mostly refers to test-scale experiments and rarely to real systems. In addition, they are typically given for fine-grained soils with respect to electroosmotic remediation, and not regarding the case of electrokinetics as a supporting measure to other remediation technologies. Thus, a realistic cost for an electrokinetic fence cannot be given. Parameters not only influencing the technical construction and remediation success but also the cost dimensions are:

�9 the objective of the application of electrokinetics, �9 the types of contaminants and/or species which are to be transported electrokinetically, �9 the site characteristics (accessibility, infrastructure), �9 the physico-chemical properties of the soil, �9 the amount of soil to be treated, �9 the hydraulic conditions, �9 the spacing of the electrodes, �9 the amount of information about the position of service pipes and metallic objects in the

subsurface.

In general the energy consumption is proportional to the electrolyte concentration of the pore solution, the field strength and the electrode spacing. Thus the higher the electric conductivity of the pore solution, the higher the energy consumption, and the higher the energy costs. The efficiency of the electric field strength is higher for electromigration compared to electroosmosis due to the direct force on the ions. For dimensioning an electrokinetic fence the choice of the electric gradient is determined mainly by the given hydraulic gradient, the ion migration velocity and the targeted groundwater constituents.

The dependency of the costs on the mentioned parameters is reflected by the great variation of electricity consumption rates observed in laboratory experiments. Depending on the types of contaminants and soils and the achieved remediation success they range between 10 kWh/m 3 (Probstein and Hicks, 1993) and 220-700 kW h/m 3 (Acar and Alshawabkeh, 1996). For the treatment of a chromate-polluted soil volume of 120 m 3, Haus and Czurda (2000) report energy costs for 6.4 kW h/m 3. Comprehensive cost models are given by Thornton and Shapiro (1995) and Alshawabkeh et al. (1999). The fixed capital costs include electrodes (installed), a generator, the electrical system, the monitoring system, any flushing equipment, and tanks and pumps, etc.; indirect costs include manpower, engineering services and installation costs; operating costs include sampling, analysis, landfill fees, maintenance and energy costs.

The costs for the application of an electrokinetic fence as supporting measure for PRBs will be less than the costs of a purely electrokinetic soil remediation activity. Several cost categories, like for preliminary geological and hydrogeological investigations, manpower and monitoring, will be common with the PRB system itself. Whether the installation of an electrokinetic fence to protect the longevity of PRB eventually is economically advantageous depends on the factors described above and has to be evaluated on a case-by-case basis.


G. Outlook

Precipitation and secondary minerals formation causing clogging and surface coating within the reactive matrix are a major problem regarding the long-term performance of PRBs. This chapter shows that the installation of an electrokinetic fence upstream of the barrier is a promising approach to reduce these phenomena.

In principle, the electric field can be applied parallel or perpendicular to the groundwater flow direction. Laboratory tests presented in this work proved the successful electrokinetic retardation of selected groundwater constituents for an electric field parallel to the groundwater flow. Computer simulations and laboratory tests of Lageman and Pool (2001) showed promising results for an electrokinetic fence placed perpendicular to the groundwater flow. Which electrode configuration will be best has to be determined for each case by adequate laboratory tests and calculations.

A successful application of electrokinetic species transport is considerably dependent on the detailed knowledge of the physico-chemical characteristics of the pore solution. Site-specific soil conditions can limit the electrokinetic processes. Therefore, detailed soil characterisation should be undertaken in addition to the standard investigation schemes. Electrokinetic transport models should be integrated in the remediation planning to optimise the application strategy. The ideal electrode configuration will depend on the geological situation and the variability in groundwater composition. The natural variability of subsurface conditions and the complexity of the treatment method will require a preliminary feasibility study. It is recommended to conduct laboratory studies using undisturbed soil samples and original pore solution or groundwater composition to simulate site conditions.

The experiences of the last decade often show that one treatment method alone is unable to achieve a satisfactory remediation result, especially if site characteristics and contamination mix are complex. In future, the combination of various remediation technologies seems to offer most promise for the successful treatment of contaminated sites. Thus, the challenge of the environmental scientist is to combine convenient technologies in a way that they support their benefits and compensate their drawbacks.

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Haus, R., 2002. Elektrokinetische Bodensanierung. Schr. Angew. Geol. Karlsruhe 65, 214p. Haus, R., Czurda, K., 1999. Electrokinetic remediation of clays. In: Kodama, H., Mermut, A.R., Torrance,

J.K. (Eds), Clays for Our Future, Ottawa, Canada, pp. 191-200. Haus, R., Czurda, K., 2000. Field scale study on in situ electroremediation, Proceedings of the Seventh

International KFK/TNO Conference on Contaminated Soil, ConSoil 2000. Telford, London, pp. 1053-1059.

Haus, R., Zorn, R., 1998. Elektrokinetische In-situ-Sanierung kontaminierter Industriestandorte. In: Czurda, K., Szabo, I. (Eds), Abfallentsorgung und Altlastensanierung, Schr. Angew. Geol. Karlsruhe, 54, pp. 93-118.

Haus, R., Zorn, R., Aldenkortt, D., 1999. Electroremediation: in-situ treatment of chromate contaminated soil. In: Yong, R.N., Thomas, H.R. (Eds), Geoenvironmental Engineering - Ground Contamination: Pollutant Management and Remediation. Telford, London, pp. 384-391.

Haus, R., Zorn, R., Czurda, K., Terfehr, S., 2002. Elektrokinetische In-situ-Sanierung. Stand der Technik, Planung, Implementierung, Arbeitskreis Innovative Erkundungs-, Sanierungs- und Uberwachungs- methoden, VoI. 7. Altlastenforum Baden-Wiirttemberg e.V., Schriftenreihe, 25p.

Hicks, R.E., Tondorf, S., 1994. Electrorestoration of metal contaminated soils. Environ. Sci. Technol. 28, 2203-2210.

Ho, S.V., Sheridan, P.W., Athmer, C.J., Heitkamp, M.A., Brackin, J.M., Weber, D., Brodsky, P.H., 1995. Integrated in situ soil remediation technology: the Lasagna process. Environ. Sci. Technol. A29, 2528-2534.

Ho, S.V., Athmer, C., Sheridan, P.W., Hughes, B.M., Orth, R., McKenzie, D., Brodsky, P.H., Shapiro, A., Thornton, R., Salvo, J., Schultz, D., Landis, R., Grifffith, R., Shoemaker, S., 1999a. The Lasagna technology for in situ soil remediation. 1. Small field test. Environ. Sci. Technol. 33, 1086-1091.

Ho, S.V., Athmer, C., Sheridan, P.W., Hughes, B.M., Orth, R., McKenzie, D., Brodsky, P.H., Shapiro, A., Sivavec, T.M., Salvo, J., Schultz, D., Landis, R., Grifffith, R., Shoemaker, S., 1999b. The Lasagna technology for in situ soil remediation. 2. Large field test. Environ. Sci. Technol. 33, 1092-1099.

Kim, J., Lee, K., 1999. Effects of electric field directions on surfactant enhanced electrokinetic remediation of diesel-contaminated sand column. J. Environ. Sci. Health A34, 863-877.

Klein, R., Schad, H., 2000. Results from a full scale funnel-and-gate system at the Beka site in Tiibingen (Germany) using zero-valent iron, Proceedings of the Seventh International KFK/TNO Conference on Contaminated Soil, ConSoil 2000. Telford, London, pp. 917-923.

Lageman, R., 1993. Electroreclamation - applications in the Netherlands. Environ. Sci. Technol. 27, 2648-2650.

Lageman, R., Pool, W., 2001. Thirteen years electro-reclamation in the Netherlands. In: Czurda, K., Haus, R., Kappeler, C., Zorn, R. (Eds), EREM 2001, Third Symposium and Status Report on Electrokinetic Remediation, Schr. Angew. Geol. Karlsruhe, 63, pp. 1/1-1/17.

MacKenzie, P.D., Homey, D.P., Sivavec, T.M., 1999. Mineral precipitation and porosity losses in granular iron columns. J. Hazard Mater. 68, 1-17.

McMahon, P.B., Dennehy, K.F., Sandstrom, M.W., 1999. Hydraulic and geochemical performance of a permeable reactive barrier containing zero-valent iron, Denver Federal Center. Ground Water 37, 396-404.

O'Hannesin, S.F., Gillham, R.W., 1998. Long-term performance of an in situ "iron wall" for remediation of VOCs. Ground Water 36, 164-170.


Probstein, R.F., 1994. Physicochemical Hydrodynamics - An Introduction, 2nd edn. Wiley, New York. Probstein, R.F., Hicks, R.E., 1993. Removal of contaminants from soil by electric fields. Science 260,

498-503.


Schnell, K., 2001. Hydrochemische Prozesse bei Weichgelinjektionen. Stoffbilanziernng, Potentielle Langzeitfolgen und Grundwassergef~ihrdungspotenziale, Schr. Angew. Geol. Karlsruhe, 61,204 p.

Steger, H., Zorn, R., Haus, R., Czurda, K., 2001. Removal of tetrachlorethylene from fine-grained soils by electrokinetic processes. In: Czurda, K., Haus, R., Kappeler, C., Zorn, R. (Eds), EREM 2001, Third Symposium and Status Report on Electrokinetic Remediation, Schr. Angew. Geol. Karlsruhe, 63, pp. 25/1-25/14.

Thornton, R.F., Shapiro, A.P., 1995. Modeling and economic analysis of in-situ remediation of Cr(VI)- contaminated soil by electromigration. In: Tedder, D.W., Pohland, F.G. (Eds), Emerging Technologies in Hazardous Waste Management V, ACS Symposium Series 607, Washington, DC, pp. 33-47.

Vogan, J.L., Focht, R.M., Clark, D.K., Graham, S.L., 1999. Performance evaluation of a permeable reactive barrier for remediation of dissolved chlorinated solvents in groundwater. J. Hazard. Mater. 68, 97-108.


Yang, G.C.C., Long, Y.-W., 1999. Removal and degradation of phenol in a saturated flow by in-situ electrokinetic remediation and Fenton-like process. J. Hazard. Mater. B69, 259-271.

Zorn, R., Haus, R., Steger, H., Czurda, K., 2000. Elektrokinetische Bodensanierung: Einsatzm6glichkeiten, Anwendungsbereiche und Erkundungsanfordernisse. Fachtagung "Elektrokinetische Verfahren - Methoden zur Altlastensanierung". LFU Bayern, Augsburg, 22 November 2000, pp. 41-52.


Chapter 9

Mecsek Ore, P~cs, Hungary case study

Mih~ily Cs6v4ri, Zsolt Berta, J6zsef Csics~ik and G4bor F6lding

A. Historical overview

Uranium mining started in southern Hungary near the city of P6cs in 1958, followed by milling in 1962. About 46 million tons of rock were extracted while the mine was operated, of which 1.4 million tons were exported to the former Soviet Union. The rest was processed on the sites of Mecsek Ore Mining Co., a predecessor of present day Mecsek Ore Environment Co. Sorting using a radiometric process produced about 19 million tons of waste.

A total of 25.8 million tons of uranium ore was chemically treated, out of which 7.2 million tons of low-grade ore were processed using alkali heap leaching, and a further 18.6 million tons were upgraded to yellow cake using the conventional acid process. The milling process resulted in 20.3 million tons of solid tailings which were placed together with approximately 32 million m 3 of process water in two tailings ponds extending over an area of 163 ha.

The former uranium mining area is situated in southern Hungary, near P6cs, a city with a population of 165,000. The schematic maps in Figures 9.1 and 9.2 show the area within Europe and Hungary and the location of the facilities.

The waste disposal facilities are concentrated in a relatively small area. Since the contaminant sources around the former mining facilities are very close to drinking water wells which provide 80% of the drinking water for the city of P6cs, there is real concern about the impact of the waste on groundwater quality. Thus one of the main objectives of the remediation work currently underway at the site is protection of the groundwater.

A general flow chart of the former ore processing and related technology is displayed in Figure 9.3. The mined rock was first sorted into three categories: waste rock, low-grade ore and ore using a radiometric measurement system.

The waste rock was transported to one of the waste rock piles (WRP), of which WRP III is the largest. Low-grade ore (with 0.01-0.03% uranium content) together with rejects from the radiometric upgrading station was treated using alkali heap leaching. The upgraded ore was then processed using conventional milling. The end product of the processes was the yellow cake (calcium diuranate). Wastes coming from mining and processing were deposited in the waste disposal areas. The contaminated mine water was cleaned to remove uranium. The company's facilities included five shafts, one mill, two heap leaching sites, two tailings ponds and three main WRPs. The mine was closed down and processing was stopped in 1997 for economic reasons (Fig. 9.4).

212 M. Cs6"vdri et al.

Figure 9.1. Location of the case study area near the city of P6cs in southern Hungary (a) and on the southern slopes of the Mecsek Mountains (b).

Mecsek Ore Environment Co.' s main activities are currently related to the remediation of the area. Most of the industrial buildings related to the ore processing have been demolished and the wastes relocated. Work is now focussing on the two tailings ponds. It is planned that the ponds will be covered in the near future, and only water treatment, groundwater restoration and monitoring will be carried out beyond the year of 2006.

Mecsek Ore, P~cs, Hungary case study 213

Figure 9.2. Uranium ore mining and processing and waste disposal facilities of the former Mecsek Ore Mining Co. near the city of P6cs.

214 M. Cs6v6ri et al.

Figure 9.3. Flow chart for the system used to process the mined rock.

Figure 9.4. View of the heap leaching area (in the foreground) and the former milling site.


Figure 9.5. Constructing the deep drainage (a) and water treatment installation (b).

Groundwater restoration started with building a contaminated water extraction system, which consists of a deep drain and a number of wells. The extracted contaminated water is treated using a lime-milk process, before it is discharged to the nearby receiver (Fig. 9.5). Approximately 0.5 million m 3 of water are removed and treated annually. The volume is expected to increase to 0.7-1 million m3/year in the future.

The use of reactive barriers (PRBs) with calcium-oxide-based reactive material for the treatment of contaminated groundwater in this area was also evaluated. For the installation of a pilot-scale experimental PRB a suitable site had to be found for field experiments on uranium removal from groundwater. This objective was one of the major tasks of the site characterisation work described in this chapter.

B. Waste characterisation

Mining and milling of uranium ore have produced large amounts of uranium-containing wastes, most of which are deposited above an aquifer near the local drinking water catchment area for the city of P~cs.

I. Characterisation of the waste rock piles

Waste rocks have been mainly disposed of in three piles: Waste Rock Pile I (WRP I). Waste rock deposited in this pile amounts to about 1.3

million tons, with an average uranium content of 70 g/t in the rock. Only a part of the leachate is collected (this has a uranium concentration of 15-20 mg/l), the rest infiltrates the subsoil under the pile and moves towards the drinking water aquifer. Monitoring wells built along the groundwater flow-path show an elevated level of uranium concentration (0.06-10 mg/l). Since heap leaching was performed near this WRP, the elevated uranium concentration may be the result of process solution that escaped from the heap leaching piles. The total dissolved solids (TDS) level in the groundwater is somewhat high (0.5-1 g/l), and the pH is nearly neutral. The uranium concentration depends on the location of the well: it reaches 10 mg/l under the waste rocks and decreases to 37-65 Ixg/l, a few hundred metres downstream from the pile.

Waste Rock Pile II (WRP II). There are about 4.4 million tons of waste in WRP II, with an average uranium content of 40 g/t. The uranium concentration in the leachate is 20-30 mg/l, and most of the leachate is collected and treated. Although a part of the

216 M. Cs6vdri et al.

Figure 9.6. Covered Waste Rock Pile II with remaining buildings of the mine.

leachate infiltrates the soil beneath the pile, it is practically impossible to carry out any further protection measures because of the unfavourable geomorphology (hills, deep valleys, fissured rock, etc.). The WRP has already been covered and it is expected that the uranium concentration in the leachate will decrease (Fig. 9.6).

Waste Rock Pile I l l (WRP Ill). The largest amount of waste rock has been disposed of in this pile (Fig. 9.7) which contains 12.3 million tons of rock with an average uranium concentration of 60 g/t. This pile is situated over the mine workings. For protection of the drinking water aquifer, mine water (containing 7 - 8 mg/l of uranium) is being pumped continuously from the mine workings. As a result of the pumping a cone of depression exists beneath the pile and a great part of the leachate from the pile converges into the mine workings. Due to the favourable hydrological situation created under this pile, with the mine workings acting as a groundwater sink, efforts have been made to collect most of the wastes here. Thus, the heap leaching wastes (7.2 million tons) have been relocated to this pile. Although most of the leachate from the waste pile is successfully collected in the mine workings, some contaminated water is escaping - mainly into the shallow aquifer in a small valley to the south of the WRP - and presents a risk to the drinking water aquifer. This was one of the reasons why the area downstream of this pile was selected as the site of the pilot-scale reactive barrier (Chapter 10).

2. Wastes from heap leaching

Low-grade ore was processed using the heap leaching technology. The heaps were arranged on two sites on which approximately 7.2 million tons of low-grade ore have been treated. After the termination of uranium production, wastes from Heap Leaching Site I


Figure 9.7. Waste Rock Pile III.

(next to the mill) have been relocated to WRP III (2.2 million tons) as the first step. Wastes from Heap Leaching Site II (the site was constructed in the vicinity of WRP I) have been relocated to WRP III in the second step. During the relocation of these wastes (approximately 5 million tons) a horizontal reactive barrier consisting of lime (1 kg/t) was used to reduce uranium migration from the wastes. A detailed description of the lime- based reactive barrier has been published (Cs6v~iri et al. 2002). The total amount of uranium in the heap leaching wastes amounted to 400 t. Photographs of the operating heap leaching are presented in Figure 9.8. The two former heap leaching sites have now been remediated.

3. Tailings from conventional milling

The two tailings ponds contain a total of 20.3 million tons of tailings (Fig. 9.9a), their average uranium concentration being 67 g/t. Because the barren process solutions were neutralised only to pH 7 - 8 (instead of shifting pH to 10.5- l 1) during the milling process, the TDS content exceeds 20 g/l in the liquid phase of the tailings. The tailings ponds were built without any proper base liner, so the process water from the ponds has infiltrated the subsoil, causing severe groundwater contamination with chemicals from the milling process (MgSO4, NaCI, etc.).

A pump-and-treat system is currently being used to treat the contaminated groundwater and thereby protect the drinking water aquifer primarily from magnesium sulphate and sodium chloride. Uranium contamination around the tailings ponds is detectable but not too high. The pump-and-treat system consists of three deep drains, some wells that collect the groundwater from the upper and deeper layers and a water treatment plant. The system has been operational since early 200 l. After discharging the free water from the tailings


Figure 9.8. Former heap leaching site during operation. The leaching solutions were distributed on top of the piles by sprinkler systems.

ponds the surface became increasingly dry (Fig. 9.9b). Remediation of tailings pond no. 2 was finished in 2003, tailings pond no.l is currently being remediated.

C. Monitoring

Mecsek Ore Environment maintains a monitoring database. The collected data cover, among other things, radiological data and the results of analyses on water samples collected from numerous monitoring wells. Data on the uranium content and other selected groundwater constituents of groundwater samples collected in the years 1996-2000 are given in Appendices A9, B9 and C9.

The monitoring wells were constructed in pairs. One of the wells is for monitoring the upper groundwater beating layer with a depth of 5 - 7 m, the other extends down to layers at greater depth (10-20 m). Thus, the contamination can be evaluated for two aquifer levels. The geological setting is such that the two aquifers are in hydrological connection with each other. The evaluation of monitoring data will be discussed - also in respect to their potential eligibility as a field test site for a pilot-scale PRB - in the following order:

Site I: Tailings ponds area Site II: Heap Leaching Area II and WRP I Site Ill: The valley where WRP III is situated Site IV: Mill site and Heap Leaching Area I


Figure 9.9. Tailings ponds in operation (a) and after closure in August 2000 (b).

1. Tailings ponds area (Site I)

The uranium concentration in the groundwater from monitoring wells around the ponds indicates that there is some uranium contamination in the area. The background uranium concentration is in the range of a few txg/l. The actual uranium concentration in the groundwater, measured in the monitoring wells (Appendix 9A), exceeds l O0 txg/l only in wells V-0I/I, V-02/I, V-2O/I, V-2I/I, V-22/1 and V-32/1. This relatively low- uranium content is most likely due to the natural ion-exchange effect in the soil underlying the tailings ponds and to the low solubility of the uranium in the deposited tailings.

220 M. Csb'vdri et al.

The specific conductivity of the groundwater is high; the main contaminants are magnesium sulphate and sodium chloride. Figure 9.10 shows the calculated TDS distribution in the groundwater. The high TDS of the groundwater (8-12 g/l) poses a serious risk to the drinking water. A pump-and-treat system has been installed to tackle this contamination.

Figure 9.11 displays the calculated distribution of the uranium contamination around the tailings ponds. The three contamination centres indicate former temporary depositions of tailings rather than contamination from the actual tailings ponds. It is expected that the contamination will rapidly decrease in these places now that relocation of the contamination sources has been completed.

The uranium concentration data from around the tailings ponds show that the contamination is in the range of some tens of Ixg/l. However, the tailings ponds must be regarded as a potential source of uranium contamination for a long time, because the overall uranium content in the tailings exceeds 1000 t.

Figure 9.10. Distribution of total dissolved solids (TDS) in the groundwater of the tailings ponds area.

Mecsek Ore, Pdcs, Hungary case study 221

Figure 9.11. Uranium contamination in the groundwater of the tailings ponds area.

2. Heap Leaching Pile H and Waste Rock Pile l (Site II)

Approximately 5 million tons of low-grade ore was treated on Heap Leaching Pile II. Waste rocks from shaft no. I have also been deposited near this site, so that there are two sources of uranium contamination. A number of wells have been used to monitor the site

and the monitoring data are presented in Appendix 9B. As indicated, there is some uranium contamination in the groundwater, but high values

were observed only in the monitoring wells Pk-O3/a, Pk-O9/a, Pk- 12/1, Pk-32/1, Pk-33/1 and in the no-longer existing process water basin. Pk-O9/a was constructed in the WRP itself, so the high uranium concentration reflects the pore water contamination in the pile itself.

222 M. Csr"vdri et al.

Specific conductivity of the groundwater exceeds 1000 IxS/cm only in a few cases, which suggests that the water has only been slightly affected by the former process solution from heap leaching. The main pollutants are sodium and sulphate: sodium originates from the reagents, sulphate from the ore. An alkaline pH was observed in a few monitoring wells, but the average pH of the water is approximately 7.5. The water from the wells with elevated uranium concentration contains higher calcium concentrations as well, which can probably be explained by pyrite oxidation within the pile and subsequent dissolution of calcite from the ore. This process may also be responsible for a relatively high conductivity of the groundwater in these wells (> 3000 IxS/cm) and a high sulphate concentration (--~ 1500 mg/l).

3. Valley below Waste Rock Pile I l l (Site Ill)

The Zsid I Valley (Fig. 9.12) is a 300 m long, 20-30 m wide and 5-10 m deep valley connecting the southern Permian of the Mecsek Mountain with the Pannonian sediments. A part of the uranium-contaminated water from WRP III is assumed to escape through the sediments in the valley. There are two permanent monitoring wells in the valley (Hb-01/l and Hb-02/l), and five provisional boreholes (Pe-01, 02, 03, 04, 05) were installed during the hydrogeological investigation of the site.

The chemistry of water samples obtained from the monitoring wells is displayed in Appendix 9C. Figures 9.13 and 9.14 show the uranium concentration and specific conductivity, respectively, in the water samples collected from the two monitoring wells. It can be seen that the uranium concentration of the water in monitoring well Hb-01/l was between 100 and 200 lxg/l at the start of measurements (1996). Since 1998, the uranium concentration has been continuously increasing, and reached 800 lxgfl by 2000. Monitoring well Hb-02/l also shows an increased uranium level but the absolute values are lower than those in well Hb-0l/l.

The increasing uranium concentration is most likely the result of large amount of wastes deposited on WRP III during recent years, and that a part of the uranium- contaminated water has escaped from the WRP through the valley. Particular attention should be paid to this increasing uranium concentration because the groundwater converges to the Pannonian sandy sediments, which are the main formations of the local drinking water aquifer. It is therefore extremely important to prevent uranium migration in the valley.

Figure 9.14 characterises the groundwater chemistry by means of specific conductivity. The curves show that the groundwater is slightly contaminated with inorganic chemicals, because the specific conductivity is elevated (> 1000 IxS/cm). The pH of the groundwater is nearly neutral, but an elevated calcium and bicarbonate concentration can be observed. Uranium concentrations in the water samples collected from monitoring well Hb-Ol/l showed the highest values among the water samples analysed (with the exception of those obtained directly from the WRP). Other chemical species in the water are common in natural waters.

1 Named after the creek in the valley.


Figure 9.12. View of the Zsid Valley and Waste Rock Pile III.

4. Mill site and Heap Leaching Pile I (Site IV)

Uranium concentration in the groundwater of Site IV decreased rapidly between 1995 and 1999 (Fig. 9.15) due to the removal of the uranium-contaminated soil. This concentration reduction slowed down in 2000, but decreased further after the site had been capped. The groundwater exhibits an elevated pH, which is the result of the former heap leaching solutions contaminating the area.

1000 -

8 0 0 "

~ ' 600 - =1.

:~ 400

200

0 01/96 01/02

" Hb-01/1

| | | | |

01/97 01/98 01/99 01/00 01/01

Sampling date

Figure 9.13. Uranium concentration in the groundwater of Site III (monitoring wells Hb-O 1/1 and Hb-02/l).


A 2000

~ 1500

1000 o.-o e j

500

0 01196 01i97 01198 01/99 01102

-4~Hb-0111 -o-HB-02/1

ol oo 01 01 Sampling date

Figure 9.14. Specific conductivity of the groundwater of Site III (monitoring wells Hb-01/1 and Hb-02/1).

D. Site characterisation, site selection

1. Geological setting of the contaminated sites

The facilities contaminated by uranium mining and processing are situated on the southern slope of the Mecsek Mountains (Fig. 9.1), and the P6cs Basin below that is filled with young sediments. Most facilities of the former mining plant, WRP and heap leaching piles were arranged on the southern slope of the mountains, which are mainly made of Permian sandstone. This sandstone formation contains the uranium ore deposits. Another part of the facilities (the mill, the tailings ponds and one of the heap leaching piles) is located just to the south within the basin.

Starting from the bottom the bed sequence of the formations in the area of interest is as follows:

Permian rocks (K6v~ig6sz616s Sandstone Formation) Lower Triassic (Jakabhegy Sandstone Formation) Pannonian formations Pleistocene and Holocene (Quaternary)

These are described in detail below (see also Figure 9.17).

25000

20000 A

m. 15000 o') --i

10000

5000

0 t 01196 01i97 01}98 01}99 01}00

Sampling date 01i01

Figure 9.15. Uranium concentration in the groundwater of Site IV.

01/02


1.1. Permian rocks (K6v6g6sz616s Sandstone Formation)

Grey sandstone (P2z e) (underlying grey sandstone). This is the most uniform member of the upper Permian bed sequence. The thickness of the grey sandstone series is 300 m. It consists of grey, fine- and medium-grained sandstone, grey, dark grey siltstone and silty, fine-grained sandstone layers. In some places fine- and medium-grained conglomerates or pebbly sandstones, coarse-grained sandstones and even thin anthracite-bearing coal layers occur. The material of the sandstones is quartz and white feldspar; muscovite also occurs in small grain sizes. The cementing material is silica. It contains a lot of plant fossils.

Green sandstone (P2z 3) (productive sandstone). The green sandstone has been investigated in the greatest detail in the area because of its uranium ore content. The rock consists of greyish green sandstone and siltstone layers, with the sandstone being predominant. The grain size of the sandstone varies, but medium sizes occur most frequently. The material of the grains is quartz, pink feldspar and muscovite. It contains pebbles of quartz porphyry, granite, metamorphites and quartz. A siliceous, sericite- beating material generally cements the rocks, but in some places carbonates play a more important role. The ore minerals fill the space between the clastic grains. These minerals contain uranium in different oxidation states and additionally coffinite and iron, copper and lead sulphides.

Red sandstone, pebbly sandstone (Pez 4) (overlying red sandstone). This is locally called the overlying red sandstone, because it directly overlies the green, productive series. Its thickness, generally 100-120 m, is increasing towards the south-east. The transition from the green sandstone is continuous. The colour boundary is predominantly sharp, but does not follow the bedding planes in every case. The Lower Triassic main conglomerate overlies it with a discordance. The pale red rocks are conglomerate, sandstone and siltstone, forming basic cycles that are several metres thick. The thickest member of the cycle consists of sandstone beds of various grain sizes. The typical layer thickness of the sandstone is 1-20 cm, most frequently 4 -5 cm, and sometimes curved cross-bedding occurs. The material of sand-sized clastic grains is quartz and pink feldspar. The pebbles are predominated by quartz porphyry varieties, in addition by quartz, and occasionally granite and metamorphic material grains also occur. No fossils have been found here, except black, siliceous wood fragments of several centimetres in diameter.

1.2. Lower Triassic (Jakabhegy Sandstone Formation) 2

The uranium ore-bearing Permian Sandstone Formation is overlain discordantly by a thick Lower Triassic sandstone series, with a thin conglomerate (main conglomerate) at its base.

e Since the map in Fig. 9.17 uses the old lithostratigraphical codes, these are applied in the text as well.

226 M. Csrvdri et al.

The red, pebbly, often diagonally bedded sandstone formation of siliceous, haematitic matrix is divided into four parts in the area:

red sandstone, siltstone (P2 zS) pale red sandstone (P2 z7) pebbly red sandstone (P2 Z6) main conglomerate (P2z 5)

1.3. Pannonian formations

Coarse gravel sand, clayey sand (Ple). The layers of the Upper Pannonian sub-stage overlie the Permian rocks in characteristic abrasion shore facies. The material of the sand beds mainly comes from the weathering of Permian rocks. The characteristic grain material is feldspar, with a slightly smaller amount of quartz. The distal facies of Upper Pannonian formations differs from the margin by the lack of gravel beds and the rocks are almost entirely grey.

1.4. Pleistocene and Holocene (Quaternary)

Slope scree and creek deposits several metres thick represent the Quaternary sediments in the valleys of the area. The hillsides in the area are often covered by loess and loess- containing soil. The uranium ore-bearing Permian sandstone is directly overlain by the porous, often pebbly Pannonian sand at many sites in the area - as illustrated by the cross-section in Figure 9.16. Thus the groundwater in the Permian series, being under higher pressure, can flow directly into the Pannonian sediments of the P~cs Basin. This is also encouraged by the faults striking NW-SE that can be observed very frequently in the Permian formations, since these faults can conduct water. This geological- hydrogeological arrangement may lead to groundwater contamination in the Pannonian Prcs Basin.

2. The geological environment of the tailings ponds

The geological structure of the Prcs Basin forming the geological environment of the tailings ponds is well known and has been discussed in many studies. Its basement consists of the oldest pre-Palaeozoic metamorphic formations of the mountain (granite and gneiss), onto which Pannonian and Pleistocene sediments of sandy and argillaceous rocks are settled in varying thickness (from 20 to several hundred metres). A cross-section of the area of the tailings ponds is shown in Figure 9.16.

The relief of the basement varies under the tailings ponds and in their immediate vicinity where the basement of the basin of metamorphic rocks rises to the surface from a depth of a few hundred metres. Therefore the total thickness of the Pannonian sediments found under the tailings ponds is only about 20-50 m. The Pannonian sediments filling the basin primarily consist of impermeable clayey and permeable sandy sediments. The permeable sandy formations store considerable water reserves, which makes this basin a significant water supply area for the city of Prcs.


Figure 9.16. Geological cross-section of the tailings ponds area.

Important structural characteristic of the sediments in the basin is that most of the layers outcrop within a relatively short distance. This behaviour is typical around the tailings ponds. In this way the impermeable layers and the aquifers cannot be clearly distinguished from each other and, most importantly, there is no continuous separating layer beneath the tailings ponds: alternating clay-containing mottled deposits and sandy layers can be found in the area.

3. Geological structure of the southern area of shaft I (Waste Rock Pile Ill)

Shaft no. I and WRP III are located on the lower part of the southern slope of the western Mecsek Mountain. This territory is the contact zone between the uranium ore-bearing Permian sandstone series constituting the mountain and the Pannonian sandy rocks filling the neighbouring southern basin, as indicated by the geological map (Fig. 9.17).

4. Screening and ranking of the sites

The evaluation of the monitoring data obtained from the four sites described in Section C has led to the conclusion that a certain level of uranium contamination occurs almost everywhere. The level is typically 50 I~g/1 in the groundwater, but in some areas it exceeds 0.5 mg/l. These values are much greater than the natural background value in this area, which is significantly less than 10 Ixg/l.


Figure 9.17. Geological map of the area (detailed investigations carried out at Site III).

From a geological point of view the suitability of the four sites for groundwater treatment by PRB can be ranked as follows:

1. the valley where WRP III is situated (Site III) 2. the heap leaching and WRP I area (Site II) 3. the tailings ponds area (Site I) 4. the mill site (Site IV)

Site III has the advantage that its morphology and subsoil structure almost constitute a natural funnel-and-gate set-up. So if the contaminant is concentrated in the upper layers a shallow barrier can easily capture it. It should also be noted that WRP III is the biggest source of uranium contamination in the area because the waste rocks deposited in it contain more than l OOO t of uranium.

Site II also exhibits an elevated uranium concentration in the groundwater at some locations, but the contamination level of the groundwater will most likely decrease without any further action as the heap leaching wastes have already been removed from the site. Nevertheless, WRP I contains approximately l O0 t of uranium and represents a lasting source of contamination.


Site I, the tailings pond area, is currently the biggest environmental problem at the former uranium mining and processing site, primarily because of the high magnesium and sodium salt contamination in the groundwater. Therefore groundwater restoration started here with the construction of a pump-and-treat system for the removal of magnesium sulphate from the groundwater. The geological structure at the site of the tailings ponds is unfavourable to the construction of a PRB even if remarkable uranium contamination would be detected because there is no continuous sealing clay layer underlying the contaminated aquifer.

Site IV can be regarded only as a temporary problem because the contaminant source has already been relocated to WRP III. As a result further discussion will be limited to Sites II and III.

E. Detailed investigation of Sites II and III

1. Geophysical investigation

A PRB usually requires an appropriate impermeable base consisting of either clay or low-permeability rock. Though general information on the potential sites for the installation of an experimental PRB was available, additional field investigations were necessary to determine the exact position of the impermeable bedrock or clay layer at Sites II and III. Geoelectrical methods were chosen for this investigation, which have proved to be effective in similar cases (Berta et al., 2000). Electrical conductivity surveys were conducted using arrays of up to 960 electrodes, and the RESP-12 computer-controlled data acquisition unit (described below) to produce two-dimensional conductivity images of relevant sections. The conductivity survey was used to distinguish rock (low electrical conductivity) and other sediments (high electrical conductivity).

The RESP-12 system is an up-to-date instrument developed in Hungary by the former E6tvOs L6r6nd Geophysical Institute. It can be used in several different operational modes. The instrument can accommodate a maximum of 960 electrodes, which are connected to distribution boxes in groups of 60. Surveys are usually carried out using an electrode spacing that depends on the depth of interest. The measurement and data processing equipment are shown schematically in Figure 9.18, which illustrates the arrangement of the geoelectrical multi-electrodes for field tests. Figure 9.19 shows the location of the measured profiles at Sites II and III.

Figure 9.18. Arrangement of geoelectrical multi-electrodes for the field tests.

230 M. Cs6v6ri et al.

Figure 9.19. Location of the profiles for geoelectrical multi-electrode measurements on Site II (PI-P3, left) and Site III (SI-S3, fight).

1.1. Conductivity survey of Site II

The measurements were carried out in the southern part of Site II. Three cross-sections were chosen: P-1 in an approximately E - W direction, and P-2 and P-3 in approximately N - S directions (Fig. 9.19). The investigated area is near WRP I. High uranium contamination was detected in monitoring well Pk-32/1 (also shown in Fig. 9.19).

The cross-section P-1 (with I m electrode spacing) indicated low conductivity in certain places. The conductivity is higher on the western part of the site, therefore, it can be expected that there is no bedrock under the upper sediments here. These findings coincide with the results obtained by a cross-section using a 5 m electrode spacing and with the results of profile P-2. The specific conductivity profile in that direction showed that there is no low- conductive zone here. More resistive layers were detected at the southern end of section P-3.

It can be concluded from the results of the geoelectrical measurements on Site II that the bedrock exhibits a very inhomogeneous morphology. However, it seems that the vicinity of monitoring well Pk-32/1 would be best suited to a PRB field test.

1.2. Conductivity survey of Site Ill

At Site III measurements were carried out along three cross-sections (see Fig. 9.19) in the valley downstream of WRP III (Fig. 9.20). Cross-section S-2 followed the flow direction of the Zsid Creek, while the two other profiles (S-I, S-3) were located on either side of the valley. The data obtained are presented in Figure 9.21.


Figure 9.20. Zsid Valley with Waste Rock Pile III.

Figure 9.21. Specific conductivity cross-section of Site III.

232 M. Cs6vdr i et al.

The conductivity survey shows that the bedrock is very near to the surface on both sides of the valley, where the low-conductivity layer can be located at a depth of 5 - 7 m in the north. This high resistivity formation is obviously the bedrock, i.e. the Permian sandstone. The bedrock disappears at about 140 m on the S-2 profile shown in Figure 9.21, and formations of higher conductivity, the Pannonian sediments, become dominant.

This observation is in accordance with the general geology of the region that suggests that the Permian sandstone is overlain by the Pannonian sediments in this area. The results of measurements using a set-up of 1 m electrode spacing indicate that the bedrock is fractured (separate dark blocks in Fig. 9.21) in the whole area. This had been expected on the basis of the general geological appearance of the area. Nevertheless the specific conductivity shows that the bedrock can be found under the upper sediments at a depth of a few metres near monitoring well Hbl / l . From this point of view the location is suitable for constructing a reactive barrier, because the barrier can extend down to the low-permeability bedrock.

2. H y d r o l o g i c a l inves t iga t ion o f S i te I l l

Based on the results of the geophysical survey the conclusion was drawn that the hard bedrock can be found under a 6 - 7 m thick layer of valley sediments (Figure 9.22). To verify this conclusion and obtain more detailed information on the characteristics of these sediments, additional boreholes were drilled at the site. These investigations were

Legend

clave:v sand ~J.f.;~2,~ sandy clay

~./~ ct, v,:~o,,~l ........ ,,, !i;~ ii p~ ...... ~rl,,,~;orP~ Q Holocen Ipluvml, deluvtat trod pahutal se&ments;

p fi6 p fi: Jakabhegvt ,S~nclstone Formation (J't'j) P f - P ; ' hS~w~g(5,~z~16xi Sandstone Formation (*P.)

Figure 9.22. Schematic geological structure of the Zsid-valley.


undertaken in two steps: in the first step the upper layer (sediments) was investigated, followed by the bedrock in the second step.

2.1. Drillings in the sediments (0-6 m)

Five boreholes (Pe-01, 02, 03, 04, 05) were sunk on Site III (Fig. 9.23). The hydraulic characteristics were determined by installing temporary wells. The details of the wells are presented in Figure 9.24. Well Pe-O 1 targeted the uppermost layer, Pe-O2 the intermediate layer and Pe-O3 and Pe-O4 the layer that is regarded as aquifer for the purposes of the experimental PRB field test. The depth intervals where grain size characteristics indicated the presence of sand were screened in the wells:

Pe-O l 1.O- 1.6 m Pe-O2 2.3-3.0 m Pe-O3 3.2-5.0 m Pe-O4 3.4-6.2 m

The permanent monitoring well Hb-Ol/l is screened from 3.5 to 8.5 m. Figure 9.25 shows the principal cross-section of the investigated area and the position of the wells.

Figure 9.23. Drilling of the boreholes in the Zsid Valley (Site III).

234 M. Csb'v6ri et al.

Figure 9.24. New boreholes logs, Site III.

Figure 9.25. Hydraulic cross-section of Site III.

The hydraulic investigation showed that the hydraulic gradient is directed upwards. The lithological characteristics of the core samples are shown in Figure 9.26. The data suggest that three local aquifers exist above the bedrock. There are no significant differences between the three aquifers with respect to their grain size characteristics (see Table 9.1). Total U concentration in the soils as well as pH and specific


Figure 9.26. Lithological characteristics of core samples from drilling Pe-O3.

conductivity of the pore water are also shown in the table. Direct hydraulic conductivity measurements were performed in the wells Hb-01/l, Pe-03 and Pe-01. Table 9.2 shows the hydraulic characteristics obtained. The permeability of the soil at a depth of 3 .2-5 m (Pe-03) is approximately 10 -5 rn/s. The upper clayey part (Pe-01) has a substantially lower hydraulic conductivity of 10 -7 m/s. For the monitoring well Hb-01/l, the permeability proved to be 2 x 10 -5 m/s.

Water samples were taken from the wells and analysed. Some of the results obtained are presented in Table 9.3. It can be seen that the uranium concentration in the samples varied between 10 and 400 txg/l. The TDS were in the range of 0.5-1.4 g/l.

Relatively high uranium concentrations were observed in well Pe-04, which was drilled at the mouth of the valley, but water in the other wells was also contaminated to some extent. It seems that the groundwater contamination is highly affected by the Zsid Creek flowing through the valley.

2.2. Drilling in the bedrock (0-15 m)

After evaluating the above results it was decided to investigate the bedrock under the 6 - 7 m thick surface sediments also. The new tests were aimed at determining the water chemistry in the bedrock and discovering the source of the uranium pollution, specifically whether it

236 M. Cs6v6r i et al.

Table 9.1. Grain size distribution, total U content, and pH and specific conductivity of the pore water in core samples of drilling Pe-03 at various depths.

Depth (m)

0.0-0.4 0.4-0.8 1.2-1.4 1.4-1.8 2.5-2.8 2.8-3.0 3.1-3.5 3.8-4.2

Grain size (%) > 4 mm 3.84 5.06 7.30 6.38 36.20 14.34 9.44 15.42

2 - 4 mm 2.86 4.90 3.92 3.40 7.20 5.32 1.46 6.36 1 -2 mm 5.22 7.26 6.18 5.58 7.74 7.14 3.54 9.18 0 .5-1 mm 10.90 16.02 15.50 14.92 14.46 15.94 19.08 17.74 0.2-0.5 mm 18.28 20.40 22.98 21.12 14.84 21.92 27.22 22.34 0 .1-0.2 mm 8.56 8.76 8.96 8.76 5.42 8.92 9.74 8.48 0.063-0.1 mm 4.16 1.82 4.66 4.14 2.14 3.98 4.10 3.46 0.045-0.063 mm 0.56 3.92 0.92 1.08 0.64 0.78 0.68 0.96 0.032-0.045 mm 2.10 1.68 2.20 2.88 0.62 1.14 0.72 0.94 < 0.032 mm 43.52 30.18 27.38 31.74 10.74 20.52 24.02 15.12

U (mg/kg) 7 8 7 < 7 < 7 < 7 < 7 < 7

pH 8.00 7.87 7.89 7.89 7.84 8.12 8.12 8.12

Electric conduc- 1020 1210 870 870 1240 560 530 740

tivity (l~S/cm)

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 1.12 • 10 -4 2.81 X 10 -5 Cooper-Jacob steptest b 8.34 x 10 -5 2.08 x 10 -5 Theis recovery test c 7.36 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

8.1 114 22 7.1 59 18 7.5 35 14 7.1 20 10 7.3 10 6 6.9 7.0 6.8 24 6 7.0 7.0 7.4 90 22 7.3 54 24

106 171 100 62 51

75

183 I43

84 39 33 20 20

23

69 60

46 39 23 19 14

19

51 41

342 300 206 99 88

135

320 340

15 774 < I 0 604 <10 318 < I 0 229 < I 0 160

< I 0 195

< I 0 488 < I 0 439

1260 1000 500 380 250

480

1400 1040

130 38 37

< 10 22 41 53 19

165 400 160 50

2 E

1426 1.78

735 2.94 509 343 2.08 684 774 593

I664 1766 1594 1205

1285 @


Figure 9.27. Drilling and testing in the bedrock (borehole Pe-I 1).

was or iginat ing f rom the water flow through the surface sediments or if it was re la ted in

large part to the fractures of the bedrock. This issue was thought to be crucial in dec id ing whe the r to construct an exper imenta l react ive barr ier at this location. Therefore an

addi t ional boreho le Pe- l l was sunk, with layer-se lec t ive water sampl ing during the drill ing.

The fo l lowing s trat igraphy was found f rom the evaluat ion of the dri l l ing cores:

O.O-l.lO m Aleurite clay: feuillemorte, tawny. Moderately sorted, aleurite content: 20-25%. Fine sand spots in some places. Limy, compact, non-stratified without partings. Contains some sandstone gravel

l . lO- 1.60 m Sandy clay: brown, rust brown from limonite. Moderately sorted, sand content: 25- 30%, in addition, it contains lO-15% aleurite. Medium-grained, well-rounded fine quartz sand. Slightly limy, compact non-stratified without parting, containing carbonised vegetation remains

1.60-2.60 m Aleurite clay: brown with slight lilac tone. Moderately sorted, its aleurite content is 20-25%; slight amounts of fine sand. Limy, compact, non-stratified without partings. Contains some sandstone gravel

2.60-3.00 m Aleurite sand: red. Moderately sorted, aleurite content: 25-30%. Medium-grained, well-rounded fine quartz sand. The aleurite "cements" the sand. It contains small amounts of square-shaped sandstone debris. Non-stratified without partings; when wet it spreads slightly in the case and hardens after desiccation

3.00-7.00 m Aleufite sand: red, pale-red. Moderately sorted, its aleurite content is 20-30%. The aleufite "cements" the sand. It contains a small amount of square-shaped sandstone debris and small amount of medium-grained gravel. The gravel is well-rounded quartzite, rhyolite and black siltstone


7.00-7.95 m

7.95-15.00 m

Sandstone debris: mainly sandstone debris, red, dull-red in some places with brownish red stripes. The rock is hard, well cemented, cherty-bonded. Coarse- grained sandstone, which contains 25% of medium-grained fraction. The grains are well-rounded quartzite. The sandstone consists of fine gravel. The gravel is well- rounded quartzite, rhyolite and black siltstone Sandstone: solid sandstone. Dull-red, laminated stratified, in some places with uncertain inclined layers. Stratification can be determined by the variation of the clay- bonded, reddish, brownish sandstone cords and sandstone cords of different grain sizes. Layer thickness: 2-30 mm, slope 40 ~ sometimes 45~ ~ The layers are usually parallel, in some cases they move towards each other (implying cross-- bedding). The rock also contains some medium-sized gravel. These are well-rounded quartzite, rhyolite and granite. The core sample is fractured an average spacing of 2- 5 cm, and the lengths of the two longest pieces were 22 and 25 cm. Extremely fractured (fractured zone): 8.60-9.00, 10.90-11.40, 12.10-12.50, 14.60-15.00 m. Slope of the joints: 60 ~ , 70 ~ , rarely 50 ~ , 90 ~

A UKB-5OO type drilling rig was used for the drilling and a packer and pump served

for the layer-selective sampling. This required a borehole of at least HQ size in order

to accommodate an NQ pipe within it for pumping purposes. The borehole diameter was 112 m m and a BASKI-type packer was used. The system was developed for HQ boreholes by the Canadian AECL. The length of the casing liner was l m, the compression hose was

equipped with 1/2 in. fittings. Nitrogen gas was used to inflate the packer and an AGA

pressure reducer provided continuous pressure. SEBA manual and D A T A Q U A automatic water gauges, and a PERSZI peristaltic pump helped perform the hydraulic tests.

Water chemistry parameters were measured using W T W Multiline F-SET3 instruments.

Figure 9.27 shows the work being performed. Six tests were carried out at different depths

and the data are displayed in Table 9.4. Table 9.5 displays selected analytical data from five water samples taken in drilling

hole Pe-I l at different depths. It can be seen that the uranium content was l OOO txg/l

in the water at a depth between O and 7 m (i.e. in the surface sediments), while only 10 -35 ixg/l in the samples obtained by packer below 7 m (i.e. in the bedrock zone). Other water characteristics also indicate that the upper groundwater significantly differs from the water coming from the bedrock. TDS was, for example, some 1400 mg/l in the upper groundwater, but only 5 0 0 - 6 0 0 mg/l in the water samples

from the bedrock zone. After finishing the planned work, the well Pe-11 was rebuilt as a monitoring well for the

bedrock ( 7 - 1 5 m depth) (Figure 9.28 ). The well is regularly monitored, and the uranium concentration in the water is the same as it was during the test period, i.e. --~ 20 lxg/l.

Table 9.4. Sampling and hydraulic testing performed in borehole Pe-I l.

Test 1 2 3 4 5 6

Date 14.5.2001 15.5.2001 15.5.2001 16.5.2001 17.5.2001 14.6.2001 Borehole bottom (m) 7.00 9.00 9.00 11.00 12.30 13.60 Open interval (m) 0-7.00 7.00-9.00 8.00-9.00 9.30-11.00 9.30-11.00 7.00-13.60 Sampling instrument Peristaltic pump Sample number F-8725 F-8726 F-8727 F-8731 F-8732 - Hydraulic test - + + + - +

h, P 0

Table 9.5. Water chemistry of samples taken from borehole Pe-11 at different depths (see Table 9.4).

Sample Date

______ -

F-8725 15.06.2001 F-8726 16.05.2001 F-8727 16.05.2001 F-8731 17.05.2001 F-8732 17.05.2COl

pH Na (mg/l)

7.38 112 7.12 32 7.12 26 6.85 19 7.31 33

K Ca (mg/l) (mgm

~~

22 200 18 96 13 96 11 89 21 84

59 54 446 26 46 130 28 37 I48 31 27 116 28 41 120

~ ~ ~~~

1400 650 550 585 535

U (CLgn)

~-

1000 35 11 10 34

Electrical

1579 712 715 657 673


Figure 9.28. Location of the monitoring wells on Site III for site investigation.

242 M. Csrvrr i et al.

These results show that the water chemistry of the bedrock zone differs significantly and reproducibly from that of the upper sediments.

F. Conclusions

High uranium concentrations have been observed during the last few years in the monitoring well Hb-Ol/l in the Zsid Valley. The source of this contamination is WRP III with approximately 1000 t of uranium remaining in the waste rock. The uranium concentration in the groundwater downstream of the WRP is still increasing, reaching values of approximately 0 . 8 - l mg/l by the end of the year 2002. The uranium contamination may have access to the drinking water aquifer downstream through this valley.

High uranium contamination has been observed in the upper sediments (6 -7 m), while the fractured zone of the bedrock, situated below 6 - 7 m, has a completely different water characteristic with low uranium concentrations.

The valley has a complicated stratigraphy and water-permeable sandy sediment layers with a permeability of 10 -4 -10 -6 m/s can be found between clay layers. The investigated part of the valley downstream of WRP III, in the immediate vicinity of monitoring wells Pe-I 1 and Hb-Ol/l, proved to be suitable for the planned pilot-scale field experiments (see Chapter 10).

Acknowledgements

The authors would like to acknowledge the help of Dr Gyrrgy Majoros and Zolt~in M~it6 in the geological description of the sites.

References

Berta, Zs., CsicsLk, J., Kov~ics, A., Varga, M., 2000. Multi-electrode geoelectric profiling to explore the spreading of a salt contamination and to design protection at a uranium processing slurry storage in Hungary. Sixth Meeting of Environmental and Engineering Geophysics (Environmental and Engineering Geophysical Society EEGS European Section), September 3-7, 2000, Bochum, Germany.

CsrvLri, 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 Ground- water Remediation. Active and Passive Technologies. Thomas Telford, London, pp. 223-235.

Appendix 9A Selected groundwater constituents analysed in water samples collected from monitoring wells at Site I (vicinity of the tailings ponds) ~

Monitoring Date Na well number (mgm

H H H H H H H H

H/1 w1 W1 H/1 w1 W1 w1 W1 W1

U-09 U-09 U-09 U-09 u -09 U-09 U-09 U-09 ~

18.04.1996 77 26.09.1996 23.04.1997 90 28.08.1997 08.04.1998 95 05.10.1998 19.05.1999 59 20.09.1999

18.04.1996 600 26.09.1996 17.04.1997 615 21.04.1997 615 28.08.1997 08.04.1998 673 05.10.1998 19.05.1999 613 20.09. I999

18.04.1996 38 30.09.1996 03.04.1997 47 27.08.1997 08.04.1998 42 05.10.1998 03.05.1999 149 16.09.1999

<5

13

<5

<5

< 5

<5 < 5

<5

<5

10

<5

9

<5

118

180

342

65

720

620 600

600

650

148

154

I80

570

13

< 10

< 10

46

1080

1285 1225

1476

1310

54

70

74

306

333 46 1 473 636 667 635 101 20 1

147 1 1489 1490 1460 1326 1663 1630 1400 1432

163 202 214 207 269 289 826 895

4 4 4 4 52 45 33 81

151 114

5185 5485 6060 5600 5460 6672 6460 5760 5600

183 263 210 220 245 273

1490 1800

<I0 18

< I 0 21

180 244

< I 0 268

-

< I 0 580

<10 671 < l o 659

< I 0 592

< 10 628

-

< 10 372

<10 323

15 412

< l O 372

-

0.80 I .06 1.11 1.70 1.40 2.15 0.70 0.7 1

10.70 10.00 12.00 11.40 10.80 12.70 13.50 11.70 11.80

0.98 1.12 0.85 I .23 1 .so 1.60 4.7 1 5.49

~~

8.9 10.4 7.4

11.3 11.7 11.5 7.5 7.0

7.3 7.1 7.0 7.0 7.4 7.2 6.9 7.1 6.8

7.6 7.3 7.6 7.4 7.5 7 .O 7.5 6.9

2 1

< I 2

< 1 < 1

5 2

17 17

107 10 8

14 12 15 20

2 2 1

< 1 < 1

1 2

37

983 I367 1387 2505 2656 2555 870 804

1 1,877 8005 8755 8635 9075

10,290 10,700

9470 9220

1450 1283 1324 1385 1612 1658 4230 4680

0.45 1.61 1.62 2.02 1 .S6 0.75 1.60 0.87

0.25 1.04 1.04 0.32 1.83 1.13 0.64 0.74 0.86

0.43 0.59 0.49 1.24 0.47 0.93 0.46 2.23

Appendix 9A (continued) 13

- .. .

u-0911 u-0911 U-09/1 u-0911 U-09/1 U-09/1 U-09/1 U-09/1

v-01 v-01 v-0 1 v-01 v-0 1 v-0 1 v-0 1 v-0 1

v-0111 v-o1/1 v-01/1 v-0111 v-o1/1 v-o1/1 v-0111 v-o2/1

Monitoring Date Na K well number (mgfl)

~

18.04.1996 255 30.09.1996 03.04.1997 300 27.08.1997 08.04.1998 407 05.10.1998 03.05.1999 403 16.09.1999

12.04.1996 85 18.09.1996 21.04.1997 96 13.09.1997 20.04.1998 81 08.09.1998 27.04.1999 80 14.09.1999

12.04.1996 132 18.09.1996 13.09.1997 20.04.1998 123 08.09.1998 27.04.1999 157 14.09.1999 03.04.1996 128

~

930

860

1188

920

26

30

28

34

I06

138

114

I 24

450

444

463

259

14

13

16

25

85

94

112

70

(276 1312 I200 1472 1524 I640 I361 I61 1

43 37 46 42 48 46 46 39

78 89 93 97 99 96 96 80

2831 2820 2550 3615 3223 3335 3290 3600

26 1 150 152 152 163 146 164 124

426 468 539 490 516 534 479 447

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

49 1

45 1

720

662

-

195

20 1

192

214

-

488 -

600

552

45 8

7.20 7.1 10 7.30 6.7 8 6.55 7.5 8 9.15 6.8 2 8.22 6.9 8 9.45 6.6 7 8.50 6.8 9 9.16 6.8 2

0.29 8.4 2 0.42 9.7 2 0.39 8.7 1 0.40 8.3 2 0.41 8.5 1 0.56 9.0 4 0.44 8.3 < 5 0.34 8.3 3

1.02 7.9 410 1.24 7.4 253 1.40 7.3 345 1.30 7.6 475 1.39 7.4 465 1.40 7.4 378 1.31 7.5 310 1.10 7.5 33

Specific Ra226 conductivity (lo-' Bq/ml) ( W c m )

8082 6025

7750 7493 8020 7250 7720

553 533 592 554 641 623 625 600

1478 1407 1689 1673 1723 1706 1698 1337

5048

0.39 0.35 0.56

18.96 0.64 1.77 0.53 0.48

0.39 0.78 0.43 1.15 0.88 1.90 0.91 0.73

0.57 0.68 0.36 1 .oo 5.08 0.78 0.52 0.37

Appendix 9A (continued) 5

v-0211 19.09.1996 v-0211 15.04.1997 140 v-0211 16.04.1997 195 v-0211 17.04.1997 200 v-0211 07.10.1997 v-0211 20.04.1998 107 v-0211 v -021 1 v-0211

V-03 V-03 V-03 V-03 V-03 V-03 V-03 V-03

v-0311 v-0311 v-0311 v-0311 v-0311 v-0311 v-0311

21.09.1998 26.04.1999 15.09.1999

04.04.1996 19.09.1996 21.04.1997 13.09.1997 21.04.1998 09.09.1998

01

28

29

20

19.04.1999 96 07.10.1999

04.04.1996 150 19.09.1996 17.04.1997 315 17.04.1997 330 13.09.1997 20.04.1998 325 09.09.1998

142 310 310

144

172

232

2 12

252

205

192

540 580

620

Mg (mgll)

~-

107 486 420

72

58

134

106

137

87

1 99

996 924

948

k

87 97

575 415

85 74 80 83 86

200 191 191 171 200 168 163 170

236 264

1080 1020 585

1000 1064

524 570 <10

2120 <I0 2230 < I 0 440 460 <10 470 486 < I 0 416

1025 <10 824 770 <10 789 830 <10 705 615 <10 750

1105 < I 0 1330 4770 < 10 4410 < I 0 2778 4370 <10 4010

-

510 525 573

519

497

390

326

439

390

-

409

439 464

476

-

1.19 7.6 30 1.40 7.5 23 4.10 7.6 57 4.10 7.5 74 1.21 7.5 43 1.18 7.6 75 1.10 7.5 160 1.20 7.6 198 1.11 7.5 355

1.80 7.4 7 1.84 7.5 7 1.67 7.9 5 1.80 7.2 5 1.90 7.6 7 1.77 7.4 7 1.50 7.7 4 1.85 7.5 17

2.20 7.7 15 2.81 7.6 17 9.00 7.1 3 8.60 7.0 32 5.40 7.2 12 8.80 7.0 20 8.70 6.9 29

v

1384 1.06 3 1543 3745 3660 1427 1483 1482 1476 1429

2045 1863 1910 1954 2204 1897 1952 2030

2330 2522 6615 6420 4792 7504 7055

-6 1.54

a 5 1.80

0.7 1 op

0.57 9 0.79 n e 1.06 m 0.71 2

0.49 & 0.55 0.56 0.49 1.02 0.52 0.92 0.34 0.72

0.40 0.29 0.62 0.72 1.29 0.73 1.86

Appendix 9A (continued) (3

v-0311 v-0311

V-18/1 V-1811 V-1811 V- 18/l V-1811 V-1811 V-18/1

v-1911 v-1911 V-19/1 V-19/1 V-1911 V-19/1 v- 1911

v-2011 v-2011 v-2011 v-2011 v-2011 v-2011 v-2011 v-2011

19.04.1999 260 07.10.1999

18.04.1996 42 30.09.1996 03.04.1997 46 27.08.1997 26.10.1998 03.05.1999 60 20.09.1999

18.04.1996 153 30.09.1996 03.04.1997 38 27.08.1997 26.10.1998 03.05.1999 222 20.09.1999

16.04.1996 59 16.10.1996 07.05.1997 107 26.08.1997 11.05.1998 91 05.10.1998 10.05.1999 73 15.09.1999

57 1

360

324

376

670

615

570

100

152

123

108

844

156

146

190

342

333

408

80

67

74

73

986 620

443 317 376 338 390 411 394

549 957 869 792 760 727 702

64 82 78 94 74 78 57 99

4000 2300

353 477 620 580 730 875 810

1830 1837 1710 1750 1820 2133 1780

298 368 420 500 315 357 257 416

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

445

445

427 -

506

415

397 -

494

488

561

537

488

-

TDS pH U Specific Ra226 (@l) conductivity ( Bqlml) (gn)

( W c m ) ~

8.40 7.0 5.30 7.2

2.50 7.4 2.00 7.1 2.00 7.4 2.30 7.1 2.46 7.0 2.81 7.1 2.88 7.0

5.30 7.3 5.00 7.2 4.60 7.4 4.80 7.0 4.70 7.2 5.43 7.2 5.00 7.1

0.53 1.6 1.10 7.3 1.15 7.4 1.35 7.0 1.04 7.4 1.20 7.2 0.62 7.3 1.31 7.3

26 27

12 11 13 11 15 18 14

6 4 3 2 5 2

12

26 113 210 260 180 195 140 220

7170 4760

3054 1839 2270 2375 2536 2742 2615

5849 4325 4300 4285 4445 4700 4230

1031 1295 1370 1720 1370 1424 1150 1574

0.46 0.86

0.57 0.69 0.36 0.66 0.52 1.23 1.65

0.40 0.39 0.58 0.79 0.42 0.44 0.85

0.56 0.82 0.57

5 2 $*

0.92 Q, 1.25 3.

0.80 1.34 1.11

Appendix 9A (continued)

Monitoring Date Na K Ca well number (mgfl) (mg/l) (mgfl)

...

v-2 1 v-2 I v-2 1 v-2 1 v-21 v-2 1 v-2 I v-2 I

v-2111 v-2111 v-21/1 v-21/1 v-21/1 v-21/1 v-21/1 v-21/1

v-22 v-22 v-22 v-22 v-22 v-22 v-22 v-22

16.04.1996 145 24.09.1996 30.04.1997 152 26.08.1997 18.05.1998 132 2 1.09.1998 10.05.1999 99 14.09.1999

16.04.1996 100 24.09.1996 30.04. I997 9 1 26.08.1997 18.05.1998 104 2 1.09.1998 10.05.1999 81 14.09.1999

05.04.1996 80 24.09.1996 07.05.1997 89 25.08.1 997 1 1.05.1998 90 26.08.1998 13.04.1999 156 15.09.1999

10

26

174

10

156

142

174

140

60

68

74

29 I

73

61

90

1 25

125

I07

119

106

17

17

18

66

103 I06 100 107 110 112 I04 103

I45 149 122 118 110 113 80

107

74 82 80 80

101 110 353 229

412 27 378 422 -

425 (10 485 488 465 <10 540 464 519 <10 546 420

447 <I0 482 477 -

450 < I 0 494 442 475 <I0 573 707 434 <10 494 389

134 <10 363 94 -

117 < I 0 299 107 85 18 290

I34 721 < I 0 345 357

Specific conductivity ( N c m )

0.93 8.8 2 1.17 8.7 5 1.25 9.0 2 1.35 7.2 4 1.37 7.2 1 1.30 7.4 3 1.35 7.3 2 1.22 7.4 6

1.31 7.7 430 1.48 7.3 338 1.20 7.5 333 1.35 7.2 371 1.36 7.2 500 1.25 7.4 340 1.22 7.6 520 1.38 7.5 270

0.46 7.9 3 0.47 8.0 2 0.55 8.0 3 0.50 8.1 3 0.54 8.2 3 0.59 8.1 3 1.96 7.4 4 1.32 7.2 7

1447 1490 1526 1674 1592 1589 1575 1551

1646 1591 1520 1720 1596 1588 1546 1559

678 644 687 720 814 86 1

2237 1568

m Ra226 f i

"2 ( Bq/ml) 0

b a, "b

1.39 c1

1.43

1.37 oc 1.23 5 0.39 3

3

0.60 c,

E 0.46 m 0.89 4

0.70 k \<

1.19 0.89 1.30 0.44 0.24 0.44 0.69

0.71 0.38 1.26 0.73 0.67 0.37 0.60 1.76


Monitoring Date Na K . .~

well number (mgfl) (md)

v-2211 v-2211 v-2211 v-2211 v-2211 v-22/1 v-221 I v-22f 1

V-23 V-23 V-23 v-23 V-23 V-23 V-23 V-23

V-23/1 V-23/1 v-2311 V-23/1 V-23/1 V-23/1 v-2311

05.04.1996 24.09.1996 07.05.1997 25.08.1997 1 1.05.1998 26.08.1998 13.04.1999 15.09.1999

24.04.1996 04.1 1.1996 29.04.1997 30.09.1997 1 I .05.1998 16.1 0.1998 22.03.1999 04.10.1999

24.04.1 996 04.1 1.1996 29.04.1997 11.05.1998 16.10.1998 22.03.1999 04.10.1999 -~ ~~

~

110

I20

132

123

60

59

62

57

122

113 110

106

Ca (mdl)

~.

90

96

120

102

46

104

88

94

104

92 131

142

______

96

94

113

111

18

50

31

40

101

109 131

163

~

67 76 80 82

108 110 104 111

60 110 62 78 46 74 35 54

85 85 76 90 94 87 97

470 < l o 470 464 -

485 <10 461 492 545 <10 503 520 509 < 10 506 490

45 < l o 302 170 -

271 <10 421 240 61 < I 0 461

234 107 < l o 503 209

484 < I 0 461 477 -

460 < 10 436 647 < I 0 482 715 702 < l o 509 700

1.05 7.6 600 1.1 1 7.4 625 1.15 7.6 658 1.23 7.7 700 1.34 7.5 800 1.32 7.3 800 1.23 7.4 710 1.32 7.7 625

0.39 8.4 2 0.67 7.5 4 0.66 7.7 2 0.81 7.5 2 0.54 7.8 2 0.85 7.3 3 0.58 7.7 2 0.66 7.7 5

1.23 7.7 20 1.14 7.6 38 1.25 7.8 28 1.58 7.5 30 1.60 7.2 13 1.65 7.4 16 1.77 7.4 16

1346 1297 1347 1441 1675 1637 1602 I600

570 879 920

1068 790

1133 850

101 1

1378 1339 1335 1742 1798 1819 1940

0.94 0.96

0.52 0.79 0.60 1.03 1.92

1.19 0.87 2.74 0.86 1.54 0.88 0.76 4.85

1.91 0.60 1.09 0.93 0.60

2.20

% 0.92 9

$1

?-

a, 2. 2

~ -

Appendix 9A (continued) ~

Monitoring Date Na K Ca well number (mgfl) tmsfl)

V-24 V-24 V-24

V-24 V-24

V-24 V-24 V-24

v-2411 V-2411 v-2411 v-2411 v-2411 V-24/ 1 v-2411 v-2411

V-25 V-25 V-25 V-25 V-25 V-25 V-25 V-25

23.04.1996 350 24.10.1996 29.04.1997 360 25.09.1 997 27.04.1998 370 28.09.1 998 06.04.1999 315 25.10.1999

23.04.1996 33 24.10. I996 29.04.1997 46 25.09.1997 27.04.1998 170 28.09.1998 06.04.1999 85 25.10.1999

03.04.1996 126 24.09.1996 25.04.1997 140 03.09.1997 07.04.1998 1 I6 24.08.1998 21.04.1999 116 08.09.1999

460

428

467

443

86

1 I4

332

207

68

70

83

82

Mg (mg/l)

~

138

120

I28

156

109

104

290

248

31

29

34

42

~

c1 ( m d u

204 163 156 150 147 152 149 170

71 78

1 I4 311 619

74 408 322

89 96 94 90 90 94 89 94

so4 (msfl)

2196 2154 2158 2069 2143 2155 2305 2150

229 296 357 584

1077 229 850 594

140 159 113 128 106 122 117 97

TDS pH

(g/l)

Specific Ra”‘ conductivity Bqlml) (Wcm)

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

< 10

214

229

220

250

-

403

354

473

555

-

464

488

525

522

-

3.70 7.0 3.35 8.1 3.50 7.7 3.37 8.0 3.50 7.4 3.50 7.4 3.60 7.4 3.63 7.5

1.00 8.0 0.85 7.9 1.05 8.0 1.85 7.6 3.60 7.3 1.05 7.7 2.50 7.5 2.22 7.5

0.71 7.6 0.73 7.5 0.66 7.6 0.72 7.4 0.77 7.6 0.69 7.3 0.71 7.6 0.73 7.5

3 3 3 2 2 3

< I 15

13 14 18 20 26 24 28 42

6 3 3 2 2 2 2 5

433 I 0.38 3100 0.90 3270 0.70 3474 0.32 3500 0.36 3460 0.66 3400 0.48 3490 0.48

1097 0.29 1030 1.42 1240 0.40 2380 0.58 3550 0.91 1199 0.64 2580 0.73 2314 0.68

979 0.49 970 0.40

1060 1.86 1034 0.86 1076 0.37 1098 1.43 1042 0.41 1042 0.75


Monitoring Date Na K well number (mgfl) (mgll)

v-2511 v-2511 v-2511 v-2511 V-2511 v-2511 v-2511 v-2511

V-29 V-29 V-29 V-29 V-29 V-29 V-29 V-29

v-2911 v-291 I V-29/1 v-2911 v-2911 V-29/1 V-29/ 1 V-2911

03.04. I996 24.09.1996 25.04.1997 03.09.1997 07.04.1998 24.08.1998 21.04.1 999 08.09.1999

24.04.1996 03.1 0.1996 13.05.1997 25.08.1997 18.05.1998 12.10.1998 13.04.1999 22.09. I999

24.04.1996 03.10.1996 13.05.1997 25.08.1997 18.05.1998 12.10.1998 13.04.1999 22.09.1999

80

92

90

92

120

119

114

100

36

33

36

19

Ca Mg (mdl) ( m a

144 98

128 115

198 119

192 124

46 14

44 <10

51 23

47 17

68 46

134 102

144 99

114 85

c1 ( m g 4

~

131 126 142 138 82

154 190 170

82 83 74 73 74 71 74 86

78 67 74 65 76 62 67 72

377 <10 512 395 -

451 <10 512 336 568 < l o 622 347 460 <10 634 400

41 (10 372 50 -

84 < I 0 366 52 49 < 10 387 35 41 (10 366 60

119 <I0 534 274 -

242 < I 0 519 486 253 < l o 497 388 212 <10 451 149

1.17 7.3 1.17 7.5 1.22 7.3 1.28 7.3 1.39 7.3 1.24 7.2 1.60 7.3 1.40 7.3

0.44 8.1 0.47 7.6 0.50 7.7 0.48 7.9 0.53 7.7 0.42 7.7 0.43 7.7 0.51 7.9

1.11 7.7 0.94 6.5 1.05 7.4 1.33 7.3 1.20 7.1 1.04 6.9 0.89 7.7 0.94 7.5

E 0

U ( F g m

~

12 13 13 10 14 15 18 18

2 3

< 1 2 2 2 1 3

10 7

10 50 14 25 12 8

1346 1380 1585 I556 1726 1683 1861 1741

716 678 710 716 76 1 731 707 726

1176 988

1254 1414 1280 1300 1069 1048

Specific Ra226 conductivity Bqlmlj (ELS/cm)

0.93 0.53 0.52 4.90 0.52 0.59 0.77 0.95

10.39 1.92 0.99 0.55 0.62 0.62 0.80 1.32

0.46 0.58 0.62 0.45 0.73 0.76

5 9 $1

1.10 a, 0.95 3.

2 %

Appendi~r 9A (continued) 3 rp ‘5

v-3211 v-3211 v-3211 V-3211 v-3211 v-3211 V-3211 V-321 I V-3211

V-3611 V-3611 V-3611 V-3611 V-3611 V-3611

X/1 X/1 X/1 X/1 X/1 XI1

XVIIU1 XVIIYI

23.04.1996 26.09.1996 15.05.1997 30.09.1997 07.10.1997 20.04.1998 21.09.1998 10.05.1999 04.10.1999

2 1 .O 1.1998 23.02.1998 29.05.1998 12.10.1998 07.04.1999 02.12.1999

20.05.1997 07.10.1997 07.05.1998 26.10.1998 25.05.1999 28.09.1999

24.10.1996 12.03.1997

Monitoring Date Na K Ca well number ( m d ) (mSn) (mg/l)

~

370

365

345

300

180 17 1 154

87 123

335

170

300

49

<5

<5

<5

<5

13 <5 <5

<5 <5

<5

<5

38

<5

5 20

500

440

364

332 312 257

220 24 1

740

352

750

220

Y rp Mg C1 SO4 COT HCO? TDS pH U Specific Ra2’6 ??

(FSIcm) 3 (md> ( m g 0 (mgll) ( m g 4 (gill (pg/l) conductivity (10 3Bqlml) 0

~ Ic

492

492

447

425

200 182 159

120 156

744

456

990

31

642 5 14 567 90

539 525 443 473 405

390 400 316 303 199 215

207 I 1489 1455 1445 3200 4798

555 537

2955 2700 2964 65 1

2580 2645 2260 2292 1890

920 1040 863 780 580 790

4675 2530 1090 2500 4370 4016

128 92

< 10

< 10

< 10

< 10

< 10 < 10

33

< 10 < 10

0

< 1 0

< 10

< 10

433

436 -

503

464

540 500 512

552 616

0

27

10

-

24

5.95 7.5 62 4.90 7.6 71 5.75 7.3 53 1.50 7.4 19 5.04 7.5 60 5.20 7.3 138 4.50 7.8 148 4.61 7.3 65 3.90 8.0 150

2.60 7.4 66 2.68 7.2 53 2.50 7.9 65 2.01 7.0 75 1.80 7.2 45 1.68 7.3 50

10.40 5.6 2 7.26 6.0 <5 6.35 6.2 30 5.70 4.6 < I 0

13.40 4.6 <5 13.74 4.6 < I 0

1.24 6.4 < I 1.00 6.9 1

Q

0.49 3

9 2

6965

5075 1 600 0.61 4495 3.09 4859 0.46

1.82 2

4320 4260 0.35 3750 1.21 w

2880 2.19 2866 0.37 2678 1.17 2424 0.42 1933 0.60 2057 0.43

4320 0.38 a 0.79 s

E

00

n

B

9060 0.64 6313 0.85 6103 0.48 5920 0.64

I 1,430 1.36 11,210 0.48

1550 2.60 1605 1.56

Appendix 9A (continued) N tn

XVIIU 1 02.09.1997 521 63 1.25 7.5 < 1 1623 2.00 XVIIYl 31.03.1998 61 < S 193 29 512 19 <10 43 1.37 7.3 < 1 1608 0.75 XVIiU1 01.09.1998 509 43 1.43 6.6 1 1547 1.56 XVIIU 1 16.03.1999 59 <S 177 32 473 62 <10 37 1.58 7.1 1 1388 0.67 X W V l 01.09.1999 471 46 1.06 6.8 15 1509

Appendix 9B Selected groundwater constituents analysed in water samples collected from monitoring wells at site II (vicinity of the former Heap Leaching Site II and Waste Rock Pile I r~

Monitoring Date Na K Ca Mg C1 804 CO3 HCO3 TDS pH U Specific Ra 226 well number (mg/l) (mg/1) (mg/l) (rag/l) (mg/1) (mg/l) (mg/l) (mg/1) (g/l) (Ixg/1) conductivity (10-4Bq/ml)

(IxS/cm)

Pk-03 11.12.1996 17 7 10 35 10 59 36 98 0.22 9.5 0.46 Pk-03 24.06.1997 16 99 0.31 8.8 < 1 385 35.75 Pk-03 09.12.1997 17 10 16 70 16 110 18 293 0.31 8.5 1 516 0.74 Pk-03 23.12.1998 12 < 5 20 54 16 66 24 265 0.35 8.9 < 1 433 0.56 Pk-03 08.12.1999 18 10 10 53 20 67 < 10 311 0.26 8.6 4 457 0.47 Pk-03 14.06.2000 14 86 0.37 8.5 7 506 0.51 Pk-03 13.12.2000 12 13 12 61 16 55 18 308 0.30 8.6 6 509 1.61

Pk-03/1 19.08.1999 36 < 5 100 48 53 100 18 412 0.67 7.7 100 824 0.67 Pk-03/1 30.09.1999 33 < 5 102 48 32 136 < 10 458 0.60 7.3 88 829 0.66 Pk-03/1 06.12.1999 16 205 0.54 7.3 50 804 0.42 Pk-03/1 27.03.2000 9 < 5 102 47 18 97 < 10 506 0.51 7.0 68 772 0.52 Pk-03/1 14.06.2000 14 215 0.52 7.2 78 763 0.71 Pk-03/1 02.10.2000 10 < 5 104 52 14 100 < 10 482 0.50 7.4 76 833 1.36 Pk-03/1 13.12.2000 12 6 79 57 14 88 < 10 479 0.55 7.2 100 764 0.93 Pk-03/1 27.02.2001 6.7 752

Pk-09/a 10.06.1996 - 53 374 - 0.87 7.4 44 1015 1.06 Pk-09/a 14.11.1996 14 < 5 148 89 57 350 < 10 445 0.95 7.7 41 1063 4.30 Pk-09/a 10.06.1997 29 324 0.95 7.2 41 1051 0.72 Pk-09/a 02.12.1997 17 < 5 142 84 51 333 < 10 448 1.01 7.4 41 1127 0.81 Pk-09/a 14.05.1998 53 312 0.96 7.2 45 1125 0.65 Pk-09/a 21.12.1998 13 < 5 130 108 34 347 < 10 397 0.83 7.0 38 1111 1.08 Pk-09/a 17.06.1999 47 291 1.03 7.3 37 1134 0.96 Pk-09/a 02.12.1999 18 < 5 136 87 54 311 < 10 458 0.91 7.4 33 1102 0.54

C~

t,,9

Appendix 9B (continued) tO

4~

Monitoring well number

Date Na

(mg/1) K (mg/1)

Ca

(mg/1) Mg (mg/1)

C1 (mg/1)

$04

(mg/1) CO3 (mg/1)

HCO3 (mg/1)

TDS (g/l)

pH U (Ixg/1)

Specific conductivity (IxS/cm)

Ra 226

(10 .4 Bq/ml)

Pk-09/a Pk-09/a

Pk-13 Pk-13 Pk-13 Pk-13 Pk-13 Pk-13 Pk-13 Pk-13 Pk-13

Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1 Pk-31/1

13.06.2000 15.11.2000

17.06.1996 12.06.1997 04.12.1997 18.06.1998 15.12.1998 17.06.1999 02.12.1999 06.06.2000 16.11.2000

30.09.1996 11.12.1996 16.02.1998 23.06.1998 23.09.1998 08.12.1998 16.03.1999 21.04.1999 30.09.1999 06.12.1999 20.03.2000 14.06.2000 28.09.2000

13

64 57 53

59

41

11 13 12

10

8

11

5

< 5

6 < 5 < 5

< 5

6

< 5 16

< 5

< 5

6

< 5

< 5

145

85 22 82

8O

85

100 101 99

80

173

73

74

85

62 63 71

96

64

32 28 26

25

50

18

20

50 50

27 20 21 27 19 22 27 29 25

25 19 28 16 14 14 30 11 11 16 12 12 11

100 278

270 372 229 189 235 193 198 171 166

204 168 154 130 109 87

369 119 91

113 76 65 76

< 1 0

< 1 0 12

< 1 0

15

< 1 0

< 1 0 0

< 1 0

< 1 0

< 1 0

< 1 0

< 1 0

448

451 342 448

488

459

226 264 250

241

329

238

229

0.81 0.85

0.73 0.70 0.69 0.44 0.80 0.73 0.72 0.55 0.55

0.43 0.53 0.48 0.46 0.41 0.43 0.92 0.44 0.35 0.33 0.41 0.41 0.39

7.3 7.4

7.6 7.8 7.6 8.0 7.2 7.6 7.7 7.4 7.8

6.6 7.1 6.8 6.8 6.5 6.5 7.2 6.8 6.8 7.0 6.5 6.7 6.6

27 37

3 2

< 5 1 2 2 8

<1 8

11

29 5 3 6

65 8 3

12 10 12 7

1081 1082

899 922 972 736 961 946 940 801 847

598

621 6O8 49O 512 944 554 450 500 459 454 517

0.59 0.68

0.32 0.69 0.60 0.46 0.90 0.63 0.73 0.48

0.93 0.49 0.44 0.96 0.79 0.77 0.48 0.56 1.26 0.42 0.55 0.33 0.97

9

t,,~~

Appendix 9B (continued)


Date Na

(mg/1) K (mg/1)

Ca

(mg/1) Mg (mg/1)

C1 (mg/1)

$04 (mg/1)

CO3 (mg/1)

HCO3 (mg/1)

TDS (g/l)

pH U (Ixg/1)

Specific conductivity (IxS/cm)

Ra 226

(10 -4 Bq/ml)

Pk-31/1

Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1 Pk-32/1

Pk-33/1 Pk-33/1 Pk-33/1 Pk-33/1 Pk-33/1

05.12.2000

30.09.1996 11.12.1996 18.03.1997 16.06.1997 02.10.1997 05.12.1997 26.02.1998 23.06.1998 28.09.1998 09.12.1998 16.03.1999 28.06.1999 29.09.1999 08.12.1999 20.03.2000 19.06.2000 28.09.2000 29.11.2000

01.12.1997 01.12.1997 16.02.1998 25.05.1998 23.09.1998

95 86 83

81

69

72

98 62

66

285 298 300

283

33 23

< 5

24

15

15

5 12

17

19 19 50

5

436 428 420

459

480

440

405 360

404

480 530 421

485

132 108 125

126

130

149

138 139

157

168 150 98

147

14

25 17 18 16 20 18 19 35 35 30

19 29 29 27 21 21 25

485 520 434 408 422

78

1555 1392 1750 1555 645

1628 1590 1682 1610 1600

1450 1458 1586 1470 1386 1389 1425

1470 1549 1383 1355 1572

< 1 0 0

< 1 0

18

< 1 0

< 1 0

< 1 0 < 1 0

< 1 0

< 1 0 < 1 0

18

< 1 0

268 258 229

293

296

336

357 314

349

433 479 409

454

0.40

2.45 2.57 2.45 2.70 2.72 2.81 2.70 2.90 2.90 2.74

2.56 2.56 2.49 2.50 2.40 2.47 2.40

3.28 3.58 3.12 2.90 3.35

6.7

7.6 7.4 7.3 7.3 7.4 7.6 6.8 7.3 6.9 6.9 7.2 7.2 7.2 7.5 6.9 7.2 7.6 8.1

7.4 7.3 7.2 7.4 7.0

4010

8290 9600

10,120 9690

10,760 12,790 12,400 12,000 10,325 11,100 10,350

5420 7370 8660 9745

7

6400 8850 5850 5710 6310

413

2203

2210 2260 2365 2438 2443 2566 2544 2520 2178 2376 2391 2387 2343 2283 2345 2285

3497 3672 3320 3460 3610

0.57

0.49 0.42 0.54 0.69 0.87 0.64 0.35 0.63 0.34 0.88 0.62 0.79 0.62 0.40 0.59 0.49 1.47 2.03

1.14 0.63 0.44 0.56 1.29

o~

r~

r~

t,~

Appendix 9B (continued)

Monitoring Date Na K Ca Mg C1 $04 CO3 HCO3 TDS pH U Specific Ra 226

well number (mg/1) (mg/1) (mg/1) (mg/1) (mg/1) (mg/1) (mg/1) (mg/1) (g/l) (Ixg/1) conductivity (10 -4 Bq/ml) (IxS/cm)

Pk-33/1 08.12.1998 422 1785 3.70 6.9 13,740 3850 Pk-33/1 26.02.1999 330 35 460 162 404 1650 < 10 421 3.60 7.2 13,960 3430 Pk-33/1 16.06.1999 408 1660 3.70 7.2 7250 3830 Pk-33/1 29.09.1999 324 1503 3.28 7.1 7820 3540 Pk-33/1 07.12.1999 314 5 381 90 276 1259 < 10 506 2.47 7.2 5420 2904 Pk-33/1 20.03.2000 268 5 475 153 340 1780 < 10 531 3.65 7.0 12,675 3740 Pk-33/1 15.06.2000 303 1504 3.23 7.3 15,220 3440 Pk-33/1 27.09.2000 235 5 427 99 285 1355 < 10 526 2.99 7.2 8730 3320 Pk-33/1 04.12.2000 290 45 524 97 308 1564 < 10 549 3.00 7.2 6020 3270

1.42 0.35 0.75 0.39 0.51 0.39 1.14 1.41 2.96

Appendix 9C Selected groundwater constituents analysed in water samples collected from monitoring wells at Site III (downstream of Waste Rock Pile III

Monitoring Date Na K Ca Mg C1 SO4 CO3 HCO3 TDS pH U Specific Ra 226 well number (mg/1) (mg/1) (mg/1) (mg/1) (mg/1) (mg/1) (mg/1) (mg/1) (g/l) (Ixg/1) conductivity (10 -4 Bq/ml)

(IxS/cm)

Hb-01/1 13.03.1996 26 < 5 82 22 19 148 < 10 232 0.41 7.8 39 545 0.87 Hb-01/1 31.05.1996 24 < 5 100 19 35 181 27 275 0.57 7.1 45 613 0.42 Hb-01/1 30.09.1996 41 18 134 40 37 196 < 10 430 0.62 7.1 158 853 0.55 Hb-01/1 11.12.1996 34 178 - 0.83 1.00 Hb-01/1 05.02.1997 293 1.07 Hb-01/1 17.02.1997 49 17 175 47 38 190 598 0.82 7.2 293 1.07 Hb-01/1 18.03.1997 40 < 5 108 19 32 167 < 10 384 0.65 7.4 163 860 0.51 Hb-01/1 16.06.1997 40 202 0.76 6.9 145 921 1.02 Hb-01/1 06.10.1997 46 21 137 12 37 240 15 418 1.02 7.0 153 957 0.40 Hb-01/1 22.10.1997 35 224 0.75 7.7 166 984 0.83 Hb-01/1 05.12.1997 39 17 139 14 35 211 <10 470 0.71 7.1 241 1009 0.89 Hb-01/1 25.02.1998 35 11 123 31 25 167 < 10 424 0.65 6.9 151 983 0.93 Hb-01/1 14.05.1998 37 15 123 26 30 172 < 10 384 0.66 7.1 130 858 1.15 Hb-01/1 17.06.1998 32 187 0.66 7.0 120 855 0.52 Hb-01/1 29.09.1998 42 15 136 44 30 243 < 10 427 0.74 6.8 160 962 0.95 Hb-01/1 04.12.1998 47 15 132 48 30 253 < 10 406 0.73 6.9 365 978 0.46 Hb-01/1 26.03.1999 35 11 102 32 18 196 < 10 375 0.60 7.0 275 832 0.36 Hb-01/1 28.06.1999 23 193 0.76 7.1 445 964 1.08 Hb-01/1 30.09.1999 57 15 140 47 34 255 < 10 421 0.90 7.1 370 1099 0.60 Hb-01/1 08.12.1999 58 15 152 47 38 279 < 10 494 0.83 7.2 605 1169 0.62 Hb-01/1 20.06.2000 44 14 140 40 32 253 < 10 412 0.75 6.8 385 1061 0.42 Hb-01/1 25.07.2000 7.0 520 1337 Hb-01/1 29.08.2000 65 20 165 60 43 330 < 10 488 1.11 7.2 475 1318 0.98 Hb-01/1 14.11.2000 68 18 169 51 39 307 < 10 445 0.89 7.0 540 1285 1.14

zz

r~

Appendix 9C (continued) ~

Monitonng well number

~~

Hb-o1/1 Hb-O1/1 Hb-Ol/l Hb-02/I Hb-02/1 Hb-02/1 Hl-02/1 Hb-02/1 Hb-02/1 Hb-02/ 1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1 Hb-02/1

11.01.2001 76 15.02.2001 15.02.2OO 1 12.03. I996 76 03.06.1996 68 25.09.1996 61 04.12.1996 02.02.1997 168 04.02.1997 19.03.1997 1 I4 16.06.1997 06.10.1997 91 06.11.1997 25.02.1998 87 17.06. I998 29.09.1998 84 16.11.1998 94 04.12.1998 89 16.03.1999 82 28.06.1999 30.09.1999 78 15.06.2000 75 02.08.2000 29.08.2000 76 29.11.2000 81

~

18

17 19 18

22

20

21

17

20 23 19 15

17 17

21 23

I70

I I6 132 116

122

126

149

151

168 152 134 136

156 152

175 177

48

43 46 48

42

49

5

45

59 51 67 52

52 54

52 60

53

71 62 55

101

85 91 71 51 58 71 58 57 62 66 so 53 57 48 54 57

374

283 235 174

265

285 83

266 340 279 266 333 295 27 1 250 255 300 313 270 327 329

~

< I 0 519

< 10 476 < I 0 488 <10 461

<10 518

< I 0 420

18 464

< 10 470

<10 482 15 500

< l o 500 < I 0 494

< I 0 500 < I 0 485

<I0 488 <I0 506

-

TDS (gAi

~~~~

1.30

0.83 0.80 0.64

1.07

1.10 0.68 0.75 1 . 1 0 1.09 1.04 1.10 1.05 1 .oo 1.02 1 .oo 1.10 1.06 1.05 1.19 1.20

PH u (PdU

~.

6.9 800 7.2 590 7.2 610 7.7 35 7.5 41 7.6 50 9.0 8.1 110 8.1 110 7.7 139 9.9 2 7.4 118 8.0 126 7.5 460 7.7 150 7.3 165 7.2 140 7.3 420 7.6 120 7.4 166 7.3 143 7.3 85 7.4 79 7.4 71 7.5 9

Specific conductivity W / c m )

1453 1745 1670 1031 1042 99 1

1319 1319 1370 904

1275 1327 1419 1332 1403 1344 1313 1184 1310 1304 1341 1426 1442 1445

~ ~ 2 2 6

( lopJ Bq/ml)

1.40

0.25 0.65 3.39 I .97 0.93 0.93 1.61 0.52 1.08 0.82 0.72 0.82 1.39 0.58 0.5 0.48 0.37 0.77

0.85

1.68

0.96 5 9

1.39 $>

2. 2 %

n, ~~~

Appendix 9C (continued)


Date Na

(mg/l) K (mg/1)

Ca

(mg/1) Mg (mg/1)

C1 (mg/1)

804 (mg/l)

CO3 (mg/l)

HCO3

(mg/l) TDS (g/l)

pH U (~g/1)

Specific conductivity (txS/cm)

Ra 226

(10 -4 Bq/ml)

Hb-02/1

Pe-01 Pe-01 Pe-02 Pe-02 Pe-03 Pe-03 Pe-03 Pe-03 Pe-04 Pe-04 Pe-04 Pe-05

26.02.2001

14.12.2000 21.02.2001 14.12.2000 21.02.2001 14.12.2000 21.02.2001 01.02.2001 07.02.2001 15.02.2001 15.02.2001 21.02.2001 21.02.2001

114 59 35 20 10 24

90 54

22 18 14 10 6 6

22 24

106 171 100 62 51 75

183 143

84 39 33 20 20 23

69 60

52

46 39 23 19 14 19

51 41

337

342 300 206

99 88

135

520 340

15 < 1 0 < 1 0 < 1 0 < 1 0 < 1 0

< 1 0 < 1 0

774 604 318 229 160 195

488 439

1.16 7.3 1433

1.26 8.1 130 1426 1.00 7.1 38 1285 0.50 7.5 37 735 0.38 7.1 < 1 0 509 0.25 7.3 22 343 0.48 6.8 19 593

6.9 41 684 7.0 53 774 7.0 165 1664 7.0 400 1766

1.40 7.4 160 1594 1.04 7.3 50 1205

1.78

2.94

2.03 r~ r r~

t,~


Chapter 10

Experimental iron barrier in P cs, Hungary

Mih~ily Cs6v~ri, J6zsef Csicszik, Gzibor F61ding and G~ibor Simoncsics

A. Introduction

The use of elemental iron (Fe ~ in permeable reactive barriers (PRBs) to remove dissolved uranium from contaminated groundwater has been proven feasible and leads to satisfactory clean-up rates (Cantrell et al., 1995; Fiedor et al., 1998). While most existing experience and results are based on laboratory column tests, there is also growing knowledge and experience available from field experiments (Phillips et al., 2000; Morrison et al., 2001; Naftz et al., 2002; Morrison, 2003). Since the behaviour of both uranium as the contaminant and elemental iron as the treatment agent is quite sensitive to the biogeochemical conditions, especially redox and pH, model experiments in the laboratory can only approximate real field conditions. Pilot field installations allow a more realistic assessment of the geochemical behaviour of the barrier material and hence of the barrier lifetime. Moreover, construction parameters, such as the appropriate homogenisation of sand/iron mixtures, and the formation of preferential pathways, can only be evaluated properly from the results of field experiments.

In this chapter the planning, installation and operation of an experimental PRB for the removal of uranium from the groundwater of a shallow aquifer is described. The field test was conducted at the former Mecsek6rc uranium ore milling and processing site near the city of P6cs in Southern Hungary, downstream of a large waste rock pile (Fig. 10.1). The site is described in detail in Chapter 9 of this volume. Based on laboratory and field tests conducted prior to designing the experimental PRB (Chapters 3-6), shredded cast iron was selected as the reactive medium.

B. Design of the permeable reactive barrier

Based on a detailed site investigation, various approaches towards the possible design of the experimental barrier were considered. These approaches included reactor-type or large column set-ups and a continuous barrier design. Due to the findings of the previous field column experiments (as described in Chapter 6) it was decided to opt for a comparatively low flow velocity within the barrier material, which could only be achieved by the continuous barrier set-up. A special advantage of this set-up was the expectation of reasonably low concentrations of dissolved iron in the barrier effluent. The experimental installation was designed with a length of 6.8 m, a thickness of 2.5 m and a depth of 3.8 m.


Figure 10.1. Location of the experimental PRB site in Zsid valley, downstream of waste rock pile III containing approximately lOOO t of uranium.

The PRB consists of two different zones: zone 1 was 50 cm thick with a low content of coarse elemental iron (12% by volume or 0.39 t/m 3, grain size 1-3 mm), and zone 2 was 1 m thick with a higher content of fine elemental iron (41% by volume or 1.28 t/m 3, grain size 0 .2-3 mm). Sand was mixed with the elemental iron to increase its volume. The total mass of elemental iron installed as reactive material was 38 t, of which 5 t was coarser material. A sketch of the design is displayed in Figure 10.2, and the technical parameters are listed in Table 10.1.

On both sides (upstream and downstream) 50-cm thick sand layers were placed to allow an even distribution of water inflow and outflow. The PRB is sealed with clay and geosynthetic clay liners at the bottom and with a geomembrane (high-density polyethylene, HDPE) at both ends and on the top, and covered over with a layer of clay. A cross-section of the experimental PRB installation is shown in Figure 10.3. The design of the PRB was submitted to the Trans-Danubian Water Authority as the responsible body, which issued the permission for the experiment.

C. Construction phase

The construction of the experimental PRB was carried out using standard construction machinery. After deciding on the exact location of the planned barrier based on the hydrogeological site investigation, a trench was excavated perpendicular to the direction of the valley. Frames, as traditionally applied in foundation engineering, were used to

Experimental iron barrier in P~cs, Hungary 263

Figure 10.2. Sketch of the design of the experimental PRB in Prcs, Hungary. Note the zones with different iron/sand mixtures in the cross-section of the barrier: low iron concentration upstream, higher iron concentration downstream.

support the walls of the excavation pit (Fig. 10.4). Significant problems occurred during construction due to large amounts of surface run-off during a period of extremely heavy rain in August 2002, which led to the complete flooding of the excavation pit on several occasions (Fig. 10.4). Moreover, due to the heavy rainfall, the work had to be carried out with a high groundwater level, and continuous dewatering of the excavation pit was therefore necessary. The two sand/iron mixtures were prepared on site using a power shovel. The different mixtures were placed manually in the protected trench, separated by wooden boards which were pulled upwards step by step as each additional layer of reactive material was placed (Fig. 10.5). After placing of the reactive material mixtures and sand layers, and installation of the monitoring wells within the barrier, the frames were lifted with a truck-mounted crane. The top of the PRB was sealed for protection with a geomembrane (HDPE), geosynthetic clay liners and a layer of compacted clay.

Table 10.1. Amount of reactive material used for construction of the PRB.

Groundwater Material Dimensions (m) Iron (ton) Iron flow through concentration barrier

Length Width Depth 0.2-2mm 1-3mm (t/m 3) (%, v/v)

Sand 6.8 0.5 3.8 Coarse Fe~

/sand 6.8 0.5 3.8 Fine Fe~

/sand 6.8 1.0 3.8 Sand 6.8 0.5 3.8

5 0.387 12

33 - 1 . 2 7 7 4 1

264 M. Csrvdri et al.

Figure 10.3. Schematic cross-section of the experimental PRB installed in the sediment' s layer of Zsid valley near P~cs, Hungary. The sediment layer above the bedrock has a total thickness of approximately 5 m.

In total, 24 monitoring wells have been installed directly in, or in the near vicinity

of the PRB. There are wells upstream, in the first iron zone, at different depths in the second iron zone, and downst ream of the barrier. At one cross-section there is a line of wells across the barrier for monitor ing precipitate formation. The wells are made of pipes of l cm diameter for groundwater sampling. Positions of the wells are shown in

Figures 10 .6-10 .8 .

Figure 10.4. Frames used for stabilisation of the excavation pit (left), flooding of the excavation pit due to heavy rainfalls during the construction phase (right).


Figure 10.5. Placement of the reactive material mixtures and the monitoring wells in the experimental PRB.

Figure 10.6. Location of monitoring wells in and around the PRB.

266 M. Cs6"vdri et al.

Figure 10. 7. View of the experimental PRB system with installed monitoring wells. The location of the PRB is indicated by a rectangle.

Groundwater flow

lo _'L_5_ , . o i *

o L 0 3 20

o 5

5 o o o 7 4 o

6

o Hb-01/1

o Pe-01

Figure 10.8. Location of the monitoring wells and points outside the barrier. The points in the barrier are shown in Figure 10.6.

Experimental iron barrier in Pdcs, Hungary 267

D. Results of operation

1. Water chemistry

The monitoring data presented in this chapter are from the period August 2002 until March 2003 when the experimental PRB system was first operational. During this 8-month period, the change in the groundwater chemistry as it passed through the barrier was observed by taking water samples from the monitoring wells. Temporal average values over this period of selected groundwater constituents in monitoring wells 1 (entrance), 8 and 10 (first zone), 9 and 11 (second zone) and 3 (exit of the barrier) are listed in Table 10.2.

The uranium concentration in the groundwater was reduced to less than 1% of its influent value by passage through the PRB (see Figure 10.9). The concentration dropped from approximately 1000 ~g/1 at the PRB inflow interface to 10 p,g/1 or even less at the PRB outflow (as indicated by the data for monitoring well PRB-3, see Table 10.2). This value is less than 20 p,g/l, which is under discussion as the limit for uranium due to its toxicity (Merkel and Sperling, 1998).

Groundwater pH was significantly increased by the elemental iron, reaching values of up to 9 - 1 0 from an initial value of 7.15 (Table 10.2). The pH was rather stable over the observation period (Fig. 10.10). In a horizontal cross-section across the barrier at a depth of 1.5 m below the top of the iron (2.3 m below the ground surface) there was a small increase in pH of the groundwater when it entered zone 1 of the barrier, where the reactive material iron is present at a lower concentration, and a larger step in the

Table 10.2. Change in water chemistry in the iron barrier (mean values of the measured monitoring data for August 2002-March 2003). Data from monitoring wells located upstream of the barrier (PRB-1), within the barrier (PRB-8 to PRB-11) and downstream of the barrier (PRB-3). Location of the monitoring wells is shown in Figure 10.6.

Inflow Zone 1: Zone 2: Outflow Change a (coarse iron/sand) (fine iron/sand)

PRB- 1 PRB-8 PRB- 10 PRB-9 PRB- 11 PRB-3

Ca mgd 167 141 136 19 21 30 - 137 Mg mgd 62 65 63 50 45 30 - 3 2 C1 mg/1 44 44 45 47 45 45 1 Fe mgd <0.1 3.89 17.0 1.05 1.91 <0.1 0 SO4 mgd 397 357 340 212 218 118 - 279 HCO3 mg~ 610 575 602 258 243 248 - 362 CO3 mgd <5 <5 <5 60 55 28 28 TDS mgd 1224 1168 947 674 680 522 -702 U ~gd 977 216 7 4 5 9 -968 pH - 7.15 7.62 7.67 9.43 9.27 9.10 1.95 Eh mV 206 13 - 122 - 84 - 94 16 - 190

a Difference between monitoring wells PRB-1 (inflow) and PRB-3 (outflow).

2 6 8 M. Csrvrr i et al.

10000

1 0 0 0 . . . . . . . . . o . . . . - o . . . . - q . . . . . 9 . . . . . . . . . . o . . . . . . . . . .

A __~. A

100 . . . . . . . . . . . . . . . . . . . . . . . . . . . . z~ . . . . . . . . . . . . . . . . . . . . .

o PRB-1 I

zx PRB-8 l

* P R B - 1 0 /

o PRB-11 /

�9 PRB-9 l

[] PRB-3 /

10

1

08 /02

/x

. . . . . . . . . _~ . . . . ~, �9 [] . . . . . . . . . . . . . . . . . . . . . . . �9 ............................. (D ..... �9 �9 [] O

o �9

. . | | ,

09/02 10/02 11/02 12/02 01/03 02/03 Date

Figure 10.9. Uranium concentration at different measuring points of the barrier system. Note the logarithmic scale on the concentration axis" location of the monitoring points see Figure 10.6.

( a ) 11 <>PRB-1 z~ P R B - 8 � 9 I o P R B - 3 * P R B - 1 0 o P R B - 1 1 I

10 . ~ ' . . ~ ~ , ~ ~ 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ~ . . . . . . . . . . .

. . . . . . . 0 . ~ _ c ~ _ _ _~_ _o_ m_ _ _- _~_ _~_ _ o . . . . _- . . . . . . . . . . . . . q"

9 ~ - ~ o ooq9 / -I- d~'o~ [] / Q.

8 ...... ~ .....................................

7- ;___:--:o;_ 6 , , , , ,

08 /02 09 /02 10 /02 11 /02 12 /02 0 1 / 0 3 0 2 / 0 3 Da te

( b ) 11 , , I i I I I I

i . . . . . . . . . J . . . . . . . . . . 10 . . . . . . . . . . . . . . . . . I , . . . . ; - I i

, I �9 9 . . . . . . . . . . . . 1 . . . . . . . . . . . . . . . . . . . . . . ~' . . . . . . . . . .

- r N t

i . . . . . . . . . . . . . . . . . . . . . . j . . . . . . . . . . u i i

t i

7 ' ' ~' . . . . . . . . . . - --2on-e-~--: ....... 2one- 2- ....... ,

I i i I

6 1 , I

0 50 100 150 2 0 0 Horizontal cross-section (cm)

Figure 10.10. (a) pH values measured at different monitoring points over the complete observation period; (b) pH 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 of monitoring wells PRB-1, 15, 16, 17, 18, 19, 20); location of the monitoring points see Figure 10.6.

(a) ._. 1800 , ,, E j . . . . . . . . . . . . . . . . . . . . j . . . . . . . . . . o 1600 . . . . . . . . . . ,

i i ~=. 1400 . . . . "- . . . . �9 . . . . . . . . . . . . . . . . . . . . ~ . . . . . . . . . . v i !

J . . . . . . . . . . . . . . . . . . . . J . . . . . . . . . . > , 1 2 0 0 . . . . . . . . . . , ..,_,

�9 - - i I �9 --- 1000 . . . . . . . . . ~ . . . . . . . . . . . . . . . . . . . . ~ . . . . . . . . . .

"~ 800 . . . . . . . . . " . . . . . . . . . . . . . . . . . . . . ~, . . . . . . . . . . " O I I �9 �9

E I �9 o 6 0 0 . . . . . . . . . ~ . . . . . . . . . . . . . . . �9 . . . . ~ . . . . . . . . . .

. . . . . . . . . . . . . . . . . . . . J . . . . . . . . . . d 400 . . . . . . . . . . r i t

2 0 0 - Z _ o n e - - - ~ . . . . . . . . . . . . . . . . . . . . l . . . . . . . . . . co~ 1 , Z o n e 2 /

0 t ,. .' I

0 50 100 150 2 0 0 Ho r i zon ta l c r o s s - s e c t i o n ( cm)

( b ) 1400 ,, ,,

1200 ' ' ' . . . . . . . . . . - t . . . . . . . . . . . . . . . . . . . . -t . . . . . . . . . �9 i I

. ~ 1000 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ~ . . . . . . . . .

800 . . . . . . . . . . J, . . . . . . . . . . . . . . . . . . . . J, . . . . . . . . . v i I

6 0 0 . . . . . . . . . . -~ . . . . . . . . . . . . . . . . . . . . 4 . . . . . . . . . I �9 �9

400 ......... ~ .................... ~ ......... t i i

2 0 0 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ~ . . . . . . . . . I Zone 1 ', Zone 2

0 .... I , ; 0 50 100 150 2 0 0

Horizontal cross-section (cm)

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.


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 ) 2 5 0

2 0 0 . . . . . . . . . . . . . . . . . . . . . . ~ . . . . . . . . . . . . . . . . . . 0�9 %,~ o o o o o 4 ~ o ~

150 . . . . -o - - ~ . . . . . . . . o ~ . . . . . . . . . . . . . . . . . . . . . . E ~ ~ o I o P R B _ 1

o ~ 100 . . . . . . t'%~ . . . . . . . . . . . . . . . . . . . . . . ~J �9176 PRB-9PRB-3 j

50 . . . . . . . . . . . . . . . . . . . . . . . . - ~ o ~ ~ ~ Jo PRB-11

0 , , , E t a , a ~, ~a ,

08/02 09/02 10/02 11/02 12/02 01/03 02/03

D a t e

( b ) 2 5 0 i Z o n e 1 Z o n e 2 I

2 0 0 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . [ . . . . . . . . .

1 5 0 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . [ . . . . . . . . . E �9 i v i

l O O . . . . . . . . . ~ . . . . . . . . . . . . . . . . . . . . t . . . . . . . . . i

50 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . , . . . . . . . . . . �9 I i

0 , I

0 5 0 1 0 0 1 5 0 2 0 0

H o r i z o n t a l c r o s s - s e c t i o n ( c m )

1 0 0

60 E

v

4 0

l o PRB-1 �9 PRB-9 J 8 0 . . . . . . . . . . . . . o . . . . . . . . u PRB-3 o PRB-111

o ~ 1 4 9 � 9 A o o . . . . z- - oO a:r- o - - ~ - 0,~ b . . . . . . . . . . . . . . . . .

oY o []~1 o

2 0 . . . . . ~ . . . . . . _o . . . . . ~ a . . . . . . . . . . . . . . . . . .

�9 o o [] oo o

0 u 'u u | n

08/02 09/02 10/02 11/02 12/02 01/03 02/03 Date

1 0 0 0

8OO

g 6 0 0

o~ 4oo o --r 200

o PRB-1 zx PRB-8 A PRB-9 [ [] PRB-3 * PRB-1O o PRB-11 . . . . . , . . . . . . ~ . . . . . . . . . . . . . . . . . . . . . . . . . . .

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

0 8 / 0 2 0 9 / 0 2 10102 1 1 / 0 2 1 2 / 0 2 0 1 / 0 3 0 2 / 0 3 D a t e

1 0 0 I Z o n e 1 Z o n e 2 '~

8 0 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . t . . . . . . . . . . ,,

6 0 . . . . u . . . . _~_ . . . . . . . . . . . . . . . . . . . . ~ . . . . . . . . . . i E ', i

~ ; 4 0 . . . . . . . . . . ', . . . . . . . . . . . . . . . . . . . . . i . . . . . . . . . . ' I i i

20 . . . . . . . . . . , . . . . . . . . . . . . . . . . . . . . . ~ . . . . . . . . , ,, i �9 0 I ,, I

0 50 1 0 0 150 2 0 0

H o r i z o n t a l c r o s s - s e c t i o n ( c m )

1 0 0 0 Z o n e 1 Z o n e 2

8 0 0 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

600 . . . . . . . . . 1 . . . . . . . . . . . . . . . . . . . . e � 9 i

u 4 0 0 . . . . . . . . . J, . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . -r" I

I �9 -4- 2 0 0 . . . . . . . . . 1 . . . . . . . . . . �9 . . . . " - - - �9 . . . . . . . . ! ! ',

0 I , I

0 50 1 0 0 1 5 0 2 0 0 H o r i z o n t a l c r o s s - s e c t i o n ( c m )

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 20O 150

~" loo 50

LU 0 -50

-100 -150 -200

I I Zone 1 ~ Zone 2 . . . . . . . . . . . . . . . . . . i . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . r . . . . . . . . . . . . . . . . .

I I

. . . . . . . . . . . . . . . . . . J . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . r . . . . . . . . . . . . . . . . .

I /

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

I I

. . . . . . . . . . . . . . . . . . i . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I- . . . . . . . . . . . . . . . . .

I I �9

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

I !

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

. . . . . . . . . . . . . . . . . . . I . . . . . . . . . . . . . . . . . . . . . . . . . . . �9 . . . . . . . . I- . . . . . . . . . . . . . . . . . I

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

I . . . . . . . �9 . . . . . . . . �9 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

I ,

0 50 100 150 21 Horizontal cross-section (cm)

Figure 10.13. Redox potential 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).


(a) 25

E 2O v t,-

.o 15 .t-.,

,4.., c

c 0

~ 5 U _

Zone l 11

--i . . . . . . . 7 I

I , 6

Zone 2 i

--o--- ~ . . . . . ! . . . . . . . . . . 10

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

. . . . . . . . . . . . . . . . o p H I - 8

- i - . . . .

0 50 100 150 200

Horizontal cross-sect ion (cm)

(b) 25

E 20 . . . . . . i . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ,,_..,

c

.__o 15 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

r -

10 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . t . )

o 5 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

O n | i i

7 8 9 10 11

pH

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, After implementation Change average of PRB a

Na mg/l 88 99 11 K mg/l 20 13.5 - 6.5 Ca mg/l 168 54 - 114 Mg mg/l 53 33 - 20 CI mg/l 43 40 - 3 $04 mg/l 361 170 - 19 l C03 mg/l < 10 < 10 - HCO3 mg/l 474 299 - 175 TDS mg/l 1030 540 - 490 pH - 7.1 7.7 0.6 U Ixg/l 718 70 - 648 Electrical ~S/cm 1496 923 - 573

conductivity

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


1 2 0 0

1 0 0 0

800

._,= 6 0 0

400

2 0 0

0

0 1 / 9 6 0 1 / 9 7 0 1 / 9 8 0 1 / 9 9 0 1 / 0 0 0 1 / 0 1 0 1 / 0 2 0 1 / 0 3

1 . 8

. . . . Imu [ ..................................................................... O 1.5 O T D S o

_1__~ . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . �9 . . . . . . . . . . �9 . . . . . . . . . 1.2

O O

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . e -m . . . . . . . . . o - t l . . . . . . . . . . . . . . . . . . . . . . . . 0 . 9 O 0 0 i j j

O O o O O O 0 � 9 1 4 9 �9 �9 l:::l 0 I-- . . . . ~__o_____o_ . . . . . . . . . o _ ~ . . . . . . . _e__m_ . . . . . . . . . . �9 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 0.6

�9 �9 0 O l �9

. . . . . . . , - . , , - - , - . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . , - - 0.3

mm mm . . . . . . . 0.0

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

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

x 23.8 x 0.06 = 2.9 x 10 -5 m3/s = 900 m3/a


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


Figure 10.16. (continued).


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 21-24 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. Vertical cross-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

. C

s6vdri et al.

~ ,-~ r~

r~

0 0 ~ ,-~

r~

! r~

0

~ ,..~ e

m

�9

0 O

r13


Figure 10.20. Simplified geological 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 10 -3 2.01 x 10 -3 1.95 x 10 -3 5.43 x 10 -3 1.11 x 10 -3 5.67 x 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.


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.



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 co- operation 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 -18 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. Location of 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 pull- apart 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-specific set-up: relief of the underlying aquitard, initial status of groundwater flow.


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

--

B9805 Meana Maxh Date" B7 Mean Max Date B9806 Mean Max Date B9808 Mean Max Date B9803 Mean Max Date

ETPH m n )

3.3 6.8 0911998

0.6 3 1111997

0.02 0.17 0512001

0.25 0.86 1111998

< 0.06 0.1 0912003

Benzene

( P m

19 29 0611998

1.15 4 0511998

0.14 3 0911998

9 17 0212000

< 1 2 1 I11999

Xylenes

( P m

435 640 0611998

36 62 041 1999

< I 3 091 1 998

23.7 43 0212000

< I ( 1

Toluene

35 50 0511989

0.5 5 Ill1997

< 1 5 0511998

6.5 18 021200 I

<1 < 1

Period 05/1998- 12/2003, main frequent respective mean value. Maximum value measured in that period. Sampling date related to the maximum value.

Ethylbenzene

CP!5!!- -

100 150 0611 998

16.6 38 1 111 997

<1 2 09/1998

25.8 47 0212000

< 1 < 1

TCE

<0.1 5.9 05/1998

<0.1 <0.1

<0.1 0.7 0511998

<0.01 5.8 05/1998

<0.1 0.15 0511998

Naphthalene

crJ-@- ~

8133 12,100 0512001

276 2100 08/2000

1.6 7.2 0511998

1 07 325 05/2001

<0.5 9.3 11/1999

~~ ~~

PAH DIN (Pgn)

~~

18 23 09/1998

12 24 0912003

0.7 I .6 021200 1

8 13 0812000

0.5 I 08/ 1999

- -

Phenols ( m m

0.22 0.34 051 199 8

0.05 0.09 03/1992

< 0.005 0.009 09/1998

0.08 0.09 1 1 /2000

< 0.005 < 0.005


Figure 11.3. Environmental site characteristics: groundwater contamination, map of former plant locations and locations of monitoring wells, hydraulic barrier and in situ adsorptive reactors.

gathered from the following investigations:

�9 historical investigation: land register, archive and permit research, multi-temporal examination of aerial photos, photogrammetric analysis of the site, investigations of existing water wells at the site and the surrounding area

�9 fifty-six investigation pits in November and December 1997 (depth maximum 5 m), of which 17 were equipped with 5 in. filter hoses before being refilled, to provide provisional monitoring wells for the determination of hydraulic heads

�9 ten core drillings (April and May 1998), equipped as 6 in. groundwater monitoring wells for sampling and water head measurements

�9 collection of approximately 300 soil samples in the unsaturated and saturated zones, and approximately 80 multi-parameter analyses of soil samples for site-specific compounds

�9 collection of water samples (two sampling series, 30 samples), analysis for general and environmental hydrochemical properties

�9 a pump test to determine the permeability and transmissivity of the sediments in the saturated zone

�9 a special investigation for PAHs and HCs �9 spatial analysis of subsurface data, reconstruction of the relief of the underlying aquitard �9 spatial analysis of the hydraulic data (reconstruction of the groundwater flow direction at

different seasonal conditions - maximum/minimum groundwater surface levels)

2. Evaluation of potential environmental hazards

Environmental assessment of the data from the site indicated a need for action. Over a large area, where industrial residues and waste had been deposited, the action thresholds,


according to limits in the relevant environmental regulations, were exceeded. Also parts of the soil in the saturated zone and the groundwater showed levels significantly above the regulatory limits according to ONORM S 2088-1 (Austrian Standards) and the Groundwater Threshold Decree (environmental law BGBI 1991).

Based on the numerous data, a site assessment of environmental hazard was performed in co-operation with experts from the Austrian Federal Environmental Agency, as required by Austrian regulations. This found that there was a potential threat to human health emanating from the site.

The site was put into the Register of Potentially Contaminated Sites (May 1998) and the Register of Contaminated Sites (May 1998) with a significant action priority. Under these requirements it was possible to apply for financial support from state funds for the remediation of the site.

3. Legal aspects

All technical measures which could affect the groundwater had to be negotiated with and approved by the relevant water authority.

D. Concept of project implementation

1. Site-specific preconditions

The aspects of the general framework that determined the system design included, for example:

�9 the site-specific hydrogeological and environmental conditions �9 the need to protect the planned artificial pond from the neighbouring groundwater

contamination (a high priority, otherwise a complete re-scheduling of the master plan for the business park would have been necessary)

�9 the proposed future use of the location as high-quality building land �9 the need to integrate implementation of the system into the infrastructure construction

work for the business park (road construction, sewer system, etc.) within the original time frame

�9 the desire to integrate the technical facilities into the business park without affecting its landscape or operation (to maintain intended image of the development)

�9 the need for acceptance by the relevant authorities (water and waste management regulators) and the public

�9 the economics of the project (both construction and running costs)

2. Evaluation of technical variants

To determine the best technical solution for improving the environmental situation at the site, and thus to enable the proposed use of the land, an extensive variant study was performed. All relevant protection and clean-up techniques were evaluated against


the background of the site-specific technical and economical conditions. The following approaches were studied:

�9 full and partial containment of the contaminated site �9 excavation of the contamination in the unsaturated and saturated zone �9 in situ and on-site microbiological treatment �9 extraction methods - soil washing in situ and on-site �9 thermal treatment �9 groundwater remediation and protection methods (conventional pump-and-treat,

innovative techniques)

Several case histories of site investigation and clean-up in Germany with similar spectra of contaminants to those found at manufactured gas plant sites were studied and the economics of the alternative remediation schemes were evaluated. Taking into account the specific technical, environmental, legal and economical constraints, several methods had to be excluded. A containment solution would have led to restrictions on the intended use of the area as building land. The financing of an extensive excavation of contaminants in both the unsaturated and saturated zones was substantially beyond the capabilities of the private investor. However, a groundwater protection and remediation strategy based on the concept of "Permeable Reactive Barriers" as presented by Starr and Cherry (1994) seemed to be applicable to the specific conditions at the site.

3. Planning of project implementation

Based on the evaluation of the site, a stepwise action plan was established to

1. remove solid contaminations (tar residuals, oil spills), contaminated foundations and underground installations (pipes, storage facilities, sewer system, etc.) from the unsaturated zone and

2. to install a groundwater protection and remediation system

The groundwater protection and remediation system was necessary to protect the planned artificial pond from the contaminant plume that otherwise would have migrated from the site towards the artificial runoff. In addition to enabling the intended development of the site, the approach to remediating the site also met the other important target of improving the environment of the surrounding area.

4. Groundwater protection

A groundwater pond of 5800m 2 was planned 40 m north of the main area of contamination as part of the landscaping concept for the business park. Its water level was planned to be 1.5 m lower than the actual groundwater level. This would have led to a nearly 180 ~ turn of the groundwater flow towards the man-made pond. This worst-case prognosis has been proven by pump tests and a hydraulic model based on the results of the hydrogeological and environmental site investigations. In order to avoid the pollution of the pond, a groundwater protection concept with different scenarios had to be established.


Figure 11.4. Basic permeable reactive barrier concept: a contaminant plume is intercepted by a reactive barrier system (left). The site adaptation as "Adsorptive Reactor and Barrier (AR&B) system": the AR&B system prevents the pollution of the artificial pond after the planned landscaping and change in groundwater flow direction (fight).

It quickly became evident that if the man-made pond was to be constructed, a cut-off wall linked to the natural barrier (high zone of underlying Tertiary aquitard) was required to protect the pond. However, without further measures the contaminated plume would still have migrated off-site, moving towards southeast. For that reason, an adaptation of the PRB concept to the site was proposed (Fig. 11.4).

E. Assessment of system- and site-specific suitabilities

1. Feasibility study

A detailed feasibility study of passive in situ systems based on the PRB concept was performed by the consulting team Niederbacher/Mapag in co-operation with the external consultants, IMES GmbH. It covered essential aspects of contaminant transport and adsorption to reactive media as well as hydraulic modelling.

1.1. Transport of organic compounds in groundwater

The main effects of the interaction of contaminants with soil matrix and groundwater are:

�9 sorption, diffusion and retardation of compounds on solid substrate of the soil matrix �9 solution, desorption and diffusion in groundwater �9 transport mechanisms in groundwater such as advection and dispersion

1.2. Water purification with activated carbon

Of the possible adsorbent media for groundwater purification, activated carbon is among the most widely used. The mechanisms of adsorption of aquatic pollutants are well known from the long-term experience in drinking water treatment. For the purification of contaminated groundwater, sorption kinetics as well as microbial degradation effects under the presence of natural and anthropogenic compounds has to be taken into account.


Research on that topic has been undertaken by a group from the University of Ttibingen, Germany, particularly with regard to a manufactured gas plant site in Karlsruhe (Grathwohl and Peschik, 1997; Schad and Grathwohl, 1998; Schad et al., 2000). This site showed a similar mix of contaminants to the site under consideration (PAHs, HCs, phenols as well as the secondary presence of chlorinated hydrocarbons). Therefore, the research results from the Karlsruhe site provided a valuable input for the Brunn am Gebirge feasibility study. The contaminant compounds found at the site in Brunn am Gebirge, and particularly the PAHs, have a high affinity to activated carbon. Therefore, activated carbon was chosen as the appropriate adsorbent.

1.3. Numeric modelling - system dimensioning

To determine the initial groundwater flow regime, and to select possible locations of the hydraulic barrier and in situ filters, numerical modelling was performed.

2. Suitability o f the permeable reactive barrier concept

The general PRB concept adapted to the site - resulting in the "AR&B system" - was evaluated and compared with conventional groundwater clean-up and protection techniques. The AR&B system was designed under site-specific geotechnical, environmental, legal and economical constraints. A feasibility study provided the validation that the intended concept was suitable to the specific site conditions. The challenge of the system design was to adapt the PRB concept to the site-specific conditions in a way which is fully integrated into the landscape of a recreation area within the business park; the remediation scheme needed to be practically invisible.

F. AR&B system implementation

Prior to constructing the AR&B system it was necessary to remove all contaminated materials from the unsaturated zone by excavation. This required the removal and disposal of approximately 80,000 ton of waste material and contaminated soil (Fig. 11.5). The excavated area was refilled with clean soil material. After this step the AR&B system was constructed. The system consisted of a hydraulic barrier, four in situ reactors with activated carbon as an adsorptive medium, the runoff system, and the monitoring system.

1. Hydraulic barrier

The PRB at this site has two main components; the hydraulic barrier and the in situ reactors (see Fig. 11.6). The hydraulic barrier is a 220 m long jet grouting wall (with an area of approximately 1000 m 2) orientated in a west-to-east direction and keyed into the tertiary shale. At the eastern end of the barrier, the tertiary shale is very shallow and forms a natural barrier. Together the jet grouting wall and the tertiary shale form an L-shaped hydraulic barrier which keeps the contaminated groundwater separated from the


Figure 11.5. Excavation of solid tar residuals.

catchment area of the artificial pond. This combination of a natural and man-made barrier induced a change of nearly 180 ~ in the local groundwater flow direction, so that it now flows towards the inlets of the in situ reactors. The water flux through the reactors is driven by the difference between the groundwater level in the contaminated area (high) and the water level of the man-made pond (low).

Figure 11.6. Scheme of the groundwater protection and remediation system "Adsorptive Reactor and Barrier (AR&B) system Brunn am Gebirge", Austria.


2. In situ adsorptive reactors and runoff system

The design of the in situ reactors had to meet the following requirements:

�9 access to the adsorptive material whenever needed �9 the possibility of permanent monitoring of the remediation system �9 a large adsorbent volume to extend the lifetime of the in situ reactors �9 high hydraulic conductivity to promote flow through the system �9 filtered entry windows to the reactor chamber positioned in the saturated zone, below

the water level of the man-made pond �9 flow driven passively by the difference in water level on the up and downstream sides of

the system

The adsorptive reactor units are located close to the barrier. To install the reactor units, four wells of 2.8 m diameter and 8 - 9 m depth were drilled. The base of the wells reached several meters into the underlying aquitard. The reactor units were made of cylindrical glass fibre reinforced synthetic material (Q 2 m, height 6 - 8 m), with filtered windows at the aquifer level. The adsorbent is activated carbon, with 1 0 - 1 2 m 3 (5-6 ton) in each reactor unit (see Fig. 11.6). Under operational conditions the contaminated water enters the reactor through filter windows, passes the reactor column and is collected at the bottom of the reactor cylinder (total flow through the whole system is approximately 0.5-2 l/s).

The outlet of each reactor passes through the barrier and is led individually to the monitoring and collecting shaft. At that spot the level of drawdown is controlled by changing the outlet level to the discharge system. The purified water is released into the drain system, which leads to the man-made pond. The outlet of this (western) pond is connected through a pipe to a second man-made pond about 200 m to the east (see Figs 11.4 and 11.6), which has its runoff to a small fiver.

Exposure of the adsorbent to an oxidising environment could cause undesirable changes of adsorptive properties and permeability (e.g. bacterial clogging). Therefore, an important construction detail of the reaction chambers is that the entrance windows are located under the minimum water level of the pond. Thus, the adsorbent cannot dry out and is never exposed to air. However, there is access into the top of the reaction chamber for visual inspection, servicing of the reactors and, if necessary, replacement of the reactive material (Fig. 11.7).

3. Adsorbent reactive material

Detailed laboratory tests investigating the behaviour of different activated carbon types with groundwater from the site have been made by Mapag GmbH with the assistance of IMES GmbH. To guarantee the necessary hydraulic conductivity, the activated carbon product "Donau Carbon CC 15" (cylindrical particles, 1.5 mm in diameter and 1.5-4 mm in length) was chosen as the adsorbent media. The selection of the type of activated carbon was based on adsorption isochrones of phenols as the least adsorptive component of the contaminants. Based on the laboratory test data, and assumptions about the expected groundwater flow and contamination load, a service life of approximately 1O-12 years was estimated for the chosen adsorptive material.


Figure 11.7. Left: Construction of a large diameter well 2.8 m in diameter and 7-9 m deep. Middle: Inside the reactor. Right: Installation of the reactor of the AR&B system.

G. System operation - hydraulics and water chemistry

The AR&B system in Brunn am Gebirge was constructed during the period of February- August 1999 and put into operation in October 1999. Since that time routine monitoring has been performed to check the hydraulic function and purification efficacy of the system. The monitoring results are showing the intended effects on groundwater protection and remediation.

1. Groundwater flow induced by the AR&B system

Since the construction was complete, groundwater from the northern (uncontaminated) part of the site now runs directly into the man-made western pond. The groundwater flow south of the hydraulic barrier is now induced to turn northwards towards the inlets of the four in situ adsorptive reactors (Figs 11.8 and 11.9). The level of drawdown is regulated by the level of the outlet in the collector shaft of the runoff system (Fig. 11.6). The outlet is a simple L-shaped pivoted pipe, connected to the drainage of the system.

The maximum drawdown can be as low as the water level of the pond. Under normal operation the water level at the entrance windows of the in situ reactors is held a few tens of centimetres above the level of the pond. As the operation of the AR&B system has caused a change in the flow direction towards the in situ reactors, an increase of the contamination content of the reactor influx could be expected - and is indeed observed - with time. Controlling the water level behind the wall can affect the water table in the catchment area of the system. Within the newly made flow system, the trough of erosional depression which is filled with coarser and more permeable sediments acts as a drainage path towards the in situ reactors G I -G3 .

The southern boundary of the catchment area shifts slightly with seasonal changes in groundwater level, but is located south of the centre of the groundwater contamination. The catchment area covers nearly 100% of the area of the groundwater contamination.


Figure 11.8. General situation of the new groundwater regime, induced by the operation of the AR&B system.

This effect minimises the migration of contaminants from the site via the distribution path of the groundwater flow. Monitoring data from the times when the groundwater level is relatively high and relatively low are shown in Figure 11.9.

2. Monitoring of groundwater chemistry and purification efficacy

Water sampling and analysis for selected organic contaminants and inorganic water chemistry is performed at selected wells (locations see Fig. 11.9) and at the inflow of each of the four reactor units in order to characterise the contaminant load (see Fig. l 1.10) entering the remediation system. A summary of the maximum concentration of selected compounds is shown in Table 11.2.

Figure 11.9. Groundwater flow induced by the operation of the AR&B system when the groundwater level is relatively high (left) and relatively low (fight).


180 Tota l B T E X

160 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . " . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . / I 140 . . . . . . . . . . . . . . . . . . . . . . . . . . . .

o~ 120 . . . . . . . . . . . . . . . . . . . . . . . . . . . . ~ ! ! ~ G 2 I . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ~ G 3 - X 100 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ----X-- G4 I l l k - -

rn 80 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ~K- "Well B7- m

o 60 b -

40

20

0 ~ 0') O') 0 0 0 0 ~ ~ ~ ~-- 04 04 04 03 03 03 03 O') O') O O O O O O O O O O O O O O O O (3) O~ O O O 0 O C) O O O O O O O O O O

, r - 04 04 04 04 04 04 04 04 04 04 04 04 04 04 04 0,4

. . . . . . 4 ' ' ' 4 - ' ' ' 4 - ' ' ' O 04 03 I'-- 0 ' ) 04 I '~ O 04 I'-.- O I'-.- 0 x - ~ O O O ~ O O ~ ~ O O ~ O O O r O

25 -r ->K / P A H s (D IN) . . . . "_ i " . . . . . . . .

2 0 . . . . . . . . . . . . . . . . . . ->1( . . . . . . . . . . . . . . . . . ; -.--- - -'-"- . . . . . . . . . . . . . .

~ " " - ;,K . . . . . .

/ ", ,: 1 Z 15 ~ . . . . . . . -,'~- - - " . - - f . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . i - -X - - G4

~ lO

5

0

1200

1000

~" 8 0 0

6 0 0

4 0 0 z

2 0 0

(3) O~ O O O O ~ r ~ ~ 04 04 04 CO 03 03 CO O) (lb O O O O O O O O O O O O O O O O (3b O~ O O O O O C:) O O O O O O O O O O

~ 04 C~l 04 04 04 C~l C~I (NI CXl ("4 04 04 (~l 04 C~I (~l , , | , , , , , , , , , i , , | | ,

O 04 03 I'-- O) 04 ~1" I'-.- O 04 ~ " I'-- O '~-- ~l" I',.- O ,,r- O O O ~'- O O ~ O O x-- O O O ' r - O

i

', ~ N a p h t h a l e n e

. . . . . . . . . . . ! i . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

' " ~ " - ~ - - G 2

---" . . . . . . -~-- . . . . . . . . . . . . . . . . . . . . . . . . . . . . I ~ G 3

- --':: . . . . . . i . . . . . . . ii . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

~ l . . . ~ n , | m ! v I v I I v I I v I A / t I I A Z ~ t l i " V t l i I W A I i

O~ O) O O O O ~ ~ ~ ~ C~ C~I 04 03 03 03 03 O~ O~ O O O O O O O O (D O O O O O O O O~ O~ O O O O O O O O O O O O O O O O

~ C~l C~l (NI (~l CXl 04 (~l 04 (~l 04 04 04 04 (NI (Xl (~l

. . . . . . 4 ' ' ' 4 . . . . . . . 0 04 r I '-. (3) 04 I'-- O ~ I'-. O , r - ~ I'-- O '~- ~ O O O ~ O O ~ ~ O O ~ O O O ~ O

Figure 11.10. Contaminant concentrations of selected compounds at the inlet of the AR&B system and well B7.

Table 11.2. Inlet and effluent concentrations of the AR&B system Brunn am Gebirge: concentrations of selected contaminants and limit concentrations set for the effluent at the collector shaft (sampling period October 1999-December 2003, mean or most frequent valueslmax values).

Parameter Unit Limit set for effluent concentration GI IN G2 IN G3 IN G4 lN Gl-G4 OUT Eftluent collector shaft

Benzene F d l 1 0.2/1.8 4.1/11 < 1/< 1 1.2/2.8 < I < 1 Ethylbenzene kg/l < I / < 1 12.8/32 3.6/16 19/68 < 1 < I Toluene k.gn 6 < 1/< 1 2.1/7.5 0.67/5.3 6.7120 < 1 < I Xylene F d l < 1/4 15.4/47 3.4/13.8 31.8169 < 1 < i BTEX pg/l 10 < 515 30/82 4/24 53/160 < S <5

PAHDIN Fgn 0.2 1.4/3.4 4.8/12 3.5/10 2.8/5.9 <0.1 (0.1 ?

I C H C 18 0.814 <0.1/2.7 < O 1/2.8 <0.1/0.3 < 1 (1 ff

Phenols mgfl 10 <0.005/0.009 0 0210.04 0.01/0.03 0.0410.07 < 0.005 < 0.005 TPH rngA 0.06 0.03/0.11 0.0210.12 0.06/0.09 0.1110.26 <0.06 < 0.06

NaphthaIene ygfl 2 3.7121 24411131 15.8183 2341693 <0.5 < 0.5 3 B

_ _ ~

9 s m T 0 3 Q

%


Measurement of the physical parameters and analysis of the water samples before and after passing through the reaction vessel ensures that the system is functioning properly. Selected relevant data have to be reported annually to the authority. Table 11.2 shows that all the important contaminants at this site are below the threshold values set by the authorities in the barrier effluent (the threshold values are also shown in Table 11.2).

Since the start-up of the system the hydraulic regime has generally changed. This is confirmed by the variations of contaminant concentrations at the inlet of the reactors. As a result of the induced change of the groundwater flow direction towards the inlet of the remediation, an increase of the contaminant content of the influent groundwater, especially of PAHs (naphthalene) and BTEX can be observed.

3. Environmental effects on site conditions

In general, the operation of the AR&B system causes a drawdown of the groundwater table compared to its initial status. This causes larger parts of the contaminated zone to be exposed to more aerated conditions, where an increase of biodegradation activity to degradable compounds can be expected. There is also a leaching effect whenever there is a precipitation event and/or rise of groundwater level, where ionic species are flushed out of the ground. This is indicated by a subsequent increase of the electric conductivity of the groundwater at the inlet of the in situ reactors after rainfall events (see Fig. 11.13).

4. Routine monitoring

An automatic monitoring system has been installed at the site (Fig. 11.11). There are water level sensors in the inlet and outlet of each reactor unit and at selected wells in the upstream area. There are resistivity/temperature sensors in the inlet and outlet of one reactor and the outlets of the other three reactors. There are also flow meters at each outlet. For general information a meteorological station (outside temperature, moisture, precipitation) is installed. All data are sampled in an interval of 15 min and wirelessly transmitted to the local data storage and transmission station. This station communicates over a distance of 27 km using an external relay station with the main station (Fig. 11.12). The supervising consultant has access to the collected data via modem for data processing.

As an example, the variations of selected relevant parameters in 2000-2002 are demonstrated in Figure 11.11. The graphs show the development of the hydraulic heads of the groundwater at the inlet and outlet of the in situ reactor (G l - G 4 ) and the electric conductivity at G3 OUT. The dataset also includes the changes of the hydraulic head in the observation well B9803 (which is upstream of the barrier) and precipitation data from the neighbouring meteorological station in V6sendorf (Fig. 11.13).

5. Advanced monitoring

The full-scale PRB system in Brunn am Gebirge allows the investigation of the reactor material and the water at different levels within the reaction vessels. During the first


Figure 11.11. Monitoring facilities at the in situ reactors and runoff system, AR&B system Brunn am Gebirge.

sampling period (August 2OOO-February 2001, with seven sampling runs) the physical and chemical water properties such as temperature, electric conductivity, redox potential, pH and oxygen content were measured in the monitoring wells of the catchment area and at the inlet, after 50 and 75% of the reactor passage, as well as at the outlet of each reactor.

Figure 11.12. Monitoring facilities - telemetry system.


Figure 11.13. Monitoring AR&B system Brunn am Gebirge 2000-2002: graphs of water levels at the reactors GI-G4 (in/out) and observation well B9803, the electric conductivity at reactor G3 out, and the precipitation at the meteorological station V6sendorf near Brunn am Gebirge.

The dataset defines the status and variations of these selected parameters at the specific sampling ports after 1 year of operation.

The dissolved oxygen content of the groundwater at the inlet of the reactors is detectable but rather low (0.1-0.8 mg/l) due to oxygen consumption in consequence of the degradation of organic contaminants in the groundwater (natural attenuation). The 02 concentrations show a seasonal variation with changes of the groundwater temperature. Chemical analyses indicate a rising contaminant content over time at the inlet of the in situ reactors as a result of the induced_groundwater flow towards and through the contamination. The observed decrease of the redox potential over time is related to this effect.

6. Permeability of the reactive material in the reactors

The chosen type of activated carbon has a defined grain size. The particles have a regular cylindrical shape of 1.5 mm diameter and 1.5-4 mm length. The resulting pore space is very uniform due to the nearly complete absence of fine-grained carbon fraction and carbon dust (the AC was washed before being filled into the reactors), a fact which is essential for the long-term performance of the reactive material.

The hydraulic effective intergranular porosity of the carbon was calculated from the field tracer test as 23%. This value matches quite well with the number 20% derived from a determination using a humid sample of reactive material (the porosity within the single grains is not part of the calculation). For the reactor G2 the height of the reactive material is l = 3.8 m. The cross-sectional area of the reactor is A = 3.14 m 2. The AC filling has a volume of 11.9 m 3. The total intergranular volume of this reactor is 2.7 m 3. The evaluation of the tracer tests indicated that the water-filled space above the reactive material of another 1.5-2 m 3 had to be drawn into calculation for every specific location. During an investigation and service run a part of the overhead volume was backfilled with fresh activated carbon so that the main part of the groundwater, which passes the entrance


windows, now flows directly into the reactive material matrix. This configuration minimizes the contact of the inflowing water with the adsorbent/free water interface, where a film of microbiological activity can clog the pore space.

As an example, at the reactor G2 a hydraulic head of h - 3 cm was observed at a flow rate of Q = 0.5 m3/h (data of January 21, 2000). From these data the permeability kf was determined:

k f - Ql _ 5.6 x 10 -3 m/s hA

In this example the calculated flow time through the reactor would be 10.4 h. This coincides with the observed time of 3 h during the tracer test at a flow rate of 1.85 m3/h.

7. Tracer tests to describe the hydraulic function o f the reactors

Much experience has been developed in recent years on the application of tracers for the investigation of natural groundwater systems. The use of water-soluble tracers in order to improve remediation techniques is often handicapped by the interaction of tracer and contaminant compounds (Nahold and H6tzl, 1992). In order to evaluate the hydraulic conductivity of the active carbon within the reaction vessels, two tracer tests (February 2001 and September 2002) were performed at each reactor (Holub, 2001). Due to the physical and chemical properties of activated carbon, no organic dye tracer can be used. Therefore, sodium chloride (NaCI) was selected as the most efficient and low-costing tracer. To check possible interactions with contaminants that could be induced by high salt concentrations, water samples were taken and analysed for contaminants during the tracer test. The tracer shows negligible interaction with the activated carbon and no significant reaction with the contaminants (certainly no significant remobilisation of contaminants was observed). The transport of the ionic NaCI tracer was logged with the conductivity sensor at the sampling port at the outlet of the reactor.

The results of the tracer tests show significant differences in flow in each of the four reactors. The residence times of the tracer peaks vary between 4 and 15 h. In Figure 11.14 the conductivity logs of two reactors (GI and G3) are compared. The tracer signal in the effluent of G I indicates a single peak breakthrough whereas the signal from G3 shows a shoulder, which is the result of two or more peaks overlapping. The reason for the observed effects can be explained by a combination of preferential flow and bulk flow.

The amount of the salt tracer was chosen in relation to the natural mineral concentration in the contaminated groundwater, which is indicated by an electric conductivity ranging from 1320 to 1900 IxS/cm (period 1999-2003: average 1450-1550 txS/cm). The salt solution was mixed into the head space so that the effect of gravity is negligible. During the pass of the tracer the electric conductivity of the effluent water of the reactors G 1 and G3 rose up to 3000 lxS/cm (conductivity at tracer maximum) showing that the quantity of salt required had been well estimated.

Table 11.3 shows the results of the tracer tests conducted in September 2002. "First tracer" indicates the time taken for the first increase of conductivity after the start of the test. This gives an estimate of the maximum velocity of transport (Va,max). "Tracer peak" indicates the elapsed time in minutes until the conductivity reached its maximum,


Figure 11.14. Tracer test, September 2002: breakthrough curves in the effluent of the reactors GI and G3 after application of the NaCI tracer.

a n d Va,me d gives an estimate of a median velocity until 50% recovery of the injected tracer. This should correspond well with a filament velocity.

Although the performance of the reactor treatment is still very good, first evidence of preferential flow through the reactive material could be detected. The cause for this could be minor clogging effects in the uppermost section of the reactive material. Traces of calcite growth were found on the surfaces of carbon grains, but mineral precipitation does not seem to have reduced the permeability significantly during the first 3 years of operation. An important outcome of these tests is that they have established a benchmark with which future performance can be compared.

8. Investigations on reactive material

As the contamination was mainly organic, activated carbon was chosen as reactive material for the reactors. The chosen reaction vessel dimensions and set-up allow sampling and investigation of any effects, which can influence the long-term behaviour of the system. In order to define the initial status during the first years of operation, quantitative data were periodically collected (e.g. the permeability of the reactive material). Beside the routine monitoring of the AR&B system according to the obligations laid down by the

Table 11.3. Results derived from the tracer tests 09/2002 in the reactors G1-G3.

Reactor Discharge First tracer v . . . . . Tracer peak Va,me d (m3/h) (min) (m/s) (min) (m/s)

Specific Specific conductivity conductivity natural water tracer max

1 0.57 376 1.77 • 10 -4 840 2 1.85 180 3.70 x 10 - 4 210 3 0.59 185 3.60 X 10 - 4 320

6.3 x 10 -5 1619 1.7 x 10 - 4 1501 7.2x 10 -5 1695

2937 2804 2558


authorities, more advanced investigations were sponsored by and performed in the course of the EU-funded PEREBAR project.

The comparison of the results of tests and analysis over time (tracer tests, investigations of the reactive material) can indicate changes in the hydraulic properties which may be caused by microbial activity, chemical precipitation of carbonates or Fe and Mn hydroxides, or due to other alteration processes.

The examination of activated carbon samples from various depths using scanning electron microscopy (SEM) revealed some small crystal phases (Fig. 11.15). Analysis of the same samples by X-ray diffraction identified only calcite as a newly formed mineral. The diffraction patterns presented in Figure l l. 16 show the comparison of fresh activated carbon with samples from the reactive material surface (O m) where the most bioactivity takes place, and a second sample from a depth of 1.5 m.

The X-ray diffractograms show that there is some precipitation of newly formed calcite over time. Besides this, no significant alteration by mineral precipitation was observed (gels or amorphous phases such as hydroxides are not detectable by X-ray diffraction). It has been widely observed that carbonate precipitation occurs when adsorptive carbon filters are used to clean water, which is heavily contaminated with mineral oil. This is because carbon dioxide is produced during microbial degradation of organic contaminants, and the carbonate produced tends to precipitate as calcite. As a consequence, the water is forced to follow preferential flow-paths, the permeability decreases, and finally the activated carbon material might experience serious clogging.

To investigate the effect of mineral precipitation (as observed in the SEM pictures), which could lead to a decrease of porosity and also adsorptive surface, AC samples from different depths within the reaction vessels were investigated using Hg porosimetry (Fig. l l. 17). To verify the results, samples were treated with different preparation methods to avoid secondary precipitation or dissolution of crystalline phases. To check the

Figure 11.15. SEM picture of an aged sample of activated carbon, precipitation of carbonate (calcite) on AC surface.


Figure 11.16. X-ray diffractograms of activated carbon: comparison of fresh activated carbon (original sample) with samples from the top (O m) and at 1.5 m depth of reactor G2 after 2 years of operation.

preparation effects, several runs were made and compared to the results of fresh AC material. The measurements show a dominating aperture range (pores of the adsorbent) of about 3000 nm. Only minor changes in the pore distribution and no significant reduction of pore space could be detected (Fig. 11.18).

An investigation of the pollutants sorbed to the activated carbon was performed (for BTEX data see Fig. 11.19). Following the downward-directed water flow through the AC, the BTEX load decreases from 120 mg/kg at the top, to about 80 mg/kg at a depth

Figure 11.17. Pore size distribution from activated carbon material (Hg porosimetry).


5O

40 - - -

30

o~ " 20

o

Original (a) 0 m depth, air- dried at 60~

0 m depth, air- dried at 105~

i

1.5 m depth, freeze-dried

Comparison of fresh AC (a) and AC after 2 years of use

Figure 11.18. Activated carbon porosity of fresh (original (a))and aged adsorbent material (reactor G2).

of 2.4 m and 19 mg/kg at 3.4 m. This investigation showed that the reactive material still has a high capacity for further contaminant adsorption; even so selective breakthrough of minor concentrations of contaminant compounds could occur due to preferential flow-paths through the AC. Also, if a reduction in the contaminant sorption capacity in the first (top) section of the reactive material occurred due to microbial growth, precipitation or contaminant overload, it could easily be replaced with a minimum of effort and costs.

BTEX-load on Activated Carbon Filter G2 (mg/kg)

0

= 0.1-0.5

(O < O {3.. 2 1.0-1.4

O

. .Q

g 2.0-2.4

llilll lilll " 0

3 0 3 4 E . - .

J

20 I I

IIIIIIIIIIIIIIIIIlIIilIIIIlH I I I I

Illlllliilllllllllll I I I I

IIIllllllllllll

40 60 80 100 120 140 I I I I I , I I I I I

�9 ' lllllllllllllllll]lllllilllllllll I I I I

I I I

I I I I

,, illiEIIflllliliiiiillllflll I I i i I I I I

}ltllllllllll}llllll]l Benzene mg/kg AC

[] Toluene mg/kg AC [] Ethylbenzene mg/kg AC [] m-Xylene mg/kg AC m o-Xylene mg/kg AC

Sample G2 1:3.01.2003 I I I I

Figure 11.19. BTEX load of activated carbon, location G2 (flow from top to bottom).


The bioactivity of omnipresent bacteria in soil and groundwater can influence a reactive or adsorptive groundwater treatment system. Activated carbon can also serve as a substrate for the bacterial fauna. To determine the actual situation in the reaction vessels, the adsorptive material was sampled at the surface of the AC material and also at a depth of 1.5 m below the surface. From these samples the basic microbiological parameters of basal respiration, substrate-induced respiration (SIR) and arginine deiminase activity were determined. The investigations were carried out by the Department of Microbiology of the University of Natural Resources and Applied Life Sciences (BOKU), Vienna, Austria.

Another important observation suggests that microbial activity preferentially develops at iron surfaces which are exposed to contaminated water. Iron-reducing bacteria can develop in large quantities within a short time. Therefore, metal parts such as filter plates, screws, etc. should be of high-quality stainless steel or coated with plastic.

The microbiological analysis on samples from the AR&B system, Brunn am Gebirge indicate that the microbial activity at a depth of 1.5 m within the reactor material is not significant, whereas at the surface of the activated carbon reactor, at the carbon/water interface, microbiological activity is evident, which can lead to the growth of a bacterial coating. This biofilm may reduce the permeability at the top of the activated carbon reactor dramatically. This problem showed up in reactor G2 at the top of the AC material where conditions were probably aerobic. As a consequence of the partly sealing microbial coating, the water level at the inlet of G2 rose during March 2000 as seen in the graph of the hydraulic head (Fig. 11.13:G2 IN). The problem was solved by the removal of the bacterial film and refilling of the top layer of activated carbon. It was important to cover the filter windows at the inlet of the system so that the main part of the influx enters the reactive material directly. As an additional measure, the free water surface within the reactor, which can vary with the water level changes, was covered with a floating air cushion film to minimise the dissolution of gaseous oxygen into the water.

Flow is not only reduced by clogging (biofouling) and precipitation of mineral phases, but also by the formation of gas within the porous material which then blocks the pore space. Gas traps installed at the top of the adsorbent material showed no evidence of gas generation, likewise no trapped gas bubbles were released by stirring the grains; so there was no evidence for the generation of a free gaseous phase. Another explanation for the absence of free gaseous phase could be that it dissolved in the groundwater.

Due to the ubiquitous presence of bacteria they are flushed into the reactor material also by the groundwater flow. Although their presence cannot be prevented totally, it is nevertheless possible to influence their activity by minimising the oxygen influx into the reactors, e.g. by coveting the free water table on top of the reactor filling.

H. Perspectives and outlook

The full-scale PRB system in Brunn am Gebirge, Austria, has operated successfully for 5 years. During the first 3 years more than 60,000 m 3 of contaminated water was treated and released to the natural water cycle. Between 20 and 40 kg of contaminants, approximately 70% of them PAHs, have been bound to the activated carbon in the four reactors of the PRB system. The flow of contaminated water through the reactors is driven


by the natural groundwater gradient across the hydraulic barrier without any energy input, and is coping well with natural variations in the local hydrologic regime.

The system design allows the function of the AR&B system to be monitored, i.e. the observation of groundwater heads and flow rates, and the sampling of groundwater on its path through the system. Advanced monitoring and testing in the course of the PEREBAR project showed that activated carbon is an appropriate adsorbent for the observed mix of organic groundwater contaminants, especially PAHs. The investigation of the adsorbent material showed alterations of the permeability, particularly at the adsorbent/water interface, caused by microbiological activity. This effect can be reduced by minor corrective measures. The most effective approach is to reduce the access of the passing groundwater to oxygen. Precipitation of minerals, mainly carbonates on the surface of the activated carbon, is currently on a microscopic scale. The amount did not increase during the observation period.

The implementation of the AR&B system, Brunn am Gebirge, Austria shows that the PRB concept can be adapted to specific environmental, geological and socio-economical site conditions. The applicability of this approach has to be evaluated by a multi- disciplinary feasibility study coveting geotechnical, hydrogeological and environmental site conditions including hydraulic modelling, investigation of groundwater chemistry and testing of reactive material under site conditions. It can also gain regulatory and public acceptance. Many of these steps are already part of the design process for conventional remediation and protection techniques.

The operation and function of the AR&B system is documented by routine monitoring, following the regulatory conditions. Routine monitoring and special investigations of barrier already in operation are useful tools in determining the factors effecting the long- term performance of a PRB system.

References

Grathwohl, P., Peschik, G., 1997. Permeable sorptive walls for treatment of hydrophobic organic contaminant plumes in groundwater. Proceedings of the International Containment Technology Conference, February 1997, St Petersburg, FL, pp. 711-717.

Holub, S., 2001. In situ Grundwassersicherung mittels Aktivkohle. Unpublished Diploma Thesis, Universit~it ftir Bodenkultur, Wien.

Nahold, M., H6tzl, H., 1992. The use of water tracers in order to improve remediation techniques. In: H6tzl, H., Werner, A. (Eds), Tracer Hydrology, Proceedings of the Sixth Symposium of Water Tracing (6th SWT), September 21-26, 1992, Karlsruhe. Balkema Publishers, Rotterdam, pp. 119-123.

Niederbacher, P., 2000. Site adapted solutions - in situ groundwater remediation "Brunn a. G.", Austria. In: Biising, J., Cortesi, P., Krejsa, P., Magiorotti, P. (Eds), ETCA Workshop on the Protection of European Water Resources: Contaminated Sites-Landfills-Sediments, Progress Review, June 21-23, 2000, Venice, Italy, pp. 253-254.

Niederbacher, P., 2001. Monitoring long-term performance of permeable reactive barriers used for the remediation of contaminated groundwater. SENSPOL Workshop "Sensing Technologies for Contaminated Sites and Groundwater", May 9-11, 2001. University of Alcal~i, Spain, pp. 162-174.

0NORM S 2088-1, 1997. Altlasten. Gef~ihrdungsabsch~itzung ftir das Schutzgut Grundwasser (Contaminated Land, Risk Assessment for Protected Groundwater). Osterreichisches Normungsin- stitut, Wien.

Schad, H., Grathwohl, P., 1998. Funnel-and-gate systems for in situ treatment of contaminated groundwater at former manufactured gas plant sites. NATO/CCMS Pilot Study, Special Session on


Treatment Walls and Permeable Reactive Barriers, February 22-28, 1998, Vienna, Austria, EPA 542- R-98-003, pp. 56-65.

Schad, H., Haist-Gulde, B., Klein, R., Maier, D., Maier, M., Schulze, B., 2000. Funnel-and-gate at the former manufactured gas plant site in Karlsruhe: sorption test results, hydraulic and technical design, construction. Proceedings of ConSoil 2000, September 18-22, 2000, Leipzig, Germany. Telford, London, pp. 951-959.

Starr, R.C., Cherry, J.A., 1994. In situ remediation of contaminated ground water: the funnel-and-gate system. Ground Water 32 (3), 465-476.

Long-term Performance of Permeable Reactive Barriers K.E. Roehl, T. Meggyes, F.-G. Simon, D.I. Stewart, editors �9 2005 Elsevier Science Ltd. All rights reserved. 311

Chapter 12

Regulatory and economic aspects

Franz-Georg Simon, Stefan Ludwig, Tam,is Meggyes, D.I. Stewart and Karl Ernst Roehl

A. Introduction

In the first 11 chapters this book has focussed primarily on the technical performance of permeable reactive barriers (PRBs), with particular emphasis on their long-term performance. However, to be routinely deployed to treat contaminated groundwater, PRBs must satisfy two other criteria: they must prove to be cost effective in comparison with other proven groundwater remediation technologies, and must be acceptable to all regulators with responsibility for the contaminated site and groundwater. This chapter reviews the issues and current progress with regards to regulatory acceptance and the cost data currently available for PRBs.

B. Regulatory aspects

1. Technology-inherent concerns

Treatment or remediation of contaminated groundwater can be very time consuming and may take time periods from years to decades due to factors such as the geological conditions in the aquifer and the hydrogeochemical behaviour of the pollutants. These factors govern the availability of the contaminants to be removed from the groundwater by a particular remediation method. Geological parameters that need to be considered are the aquifer's composition (especially the presence of fine-grained, clayey sediments with a high contaminant retention capacity), its spatial structure and its overall hydraulic permeability. The most important properties of the contaminants are their speciation and geochemical behaviour, particularly their mobility under different redox and pH conditions. Active groundwater remediation methods should be applied if the pollutant concentration is high and decontamination measures must be taken immediately. Passive remediation technologies, such as PRBs, tend to work over longer periods of time, and can be used if long treatment times are acceptable. Thus assessment of the long-term performance of PRBs requires a full understanding of the underlying processes involved. The main processes employed within PRBs are:

�9 reductive degradation of organic pollutants; �9 oxidative degradation of organic pollutants; �9 retardation and biodegradation of organic pollutants;

312 F.-G. Simon et al.

�9 sorption of organic or inorganic pollutants; �9 reduction and/or precipitation of heavy metal compounds.

The first three processes lead to a decomposition of the pollutants without them being retained within the barrier. However, the degradation processes will produce metabolites whose fate must be considered. This is particularly relevant in the case of the treatment of chlorinated hydrocarbons. The degradation of tetrachloroethylene (PCE) results in the formation of metabolites such as trichloroethylene (TCE), cis-l,2-dichloroethylene (cis-l,2-DCE) and even vinyl chloride which themselves have to be treated effectively in the barrier in order to avoid their release to the environment.

Pollutants that remain in the barrier (sorbed, precipitated or immobilised compounds) may cause various problems. They may clog the pores of the barrier, undergo chemical reactions over the long term or get desorbed or re-mobilised and create a secondary source of pollution. Several options are available to overcome these difficulties:

�9 removal and replacement of reactive material; �9 monitoring of the barrier; �9 stabilisation of the reactive material; �9 bio-mineralisation (although the appropriate method is not yet available).

If the pollutant remains within the PRB it is important to ensure that the retardation is irreversible. Figure 12.1 illustrates potential contamination histories and effects of various treatments in terms of pollutant concentration as a function of time. The groundwater pollution which occurs at some point downstream of the contaminant source when no treatment is undertaken is shown by curve 1. If the retardation by the PRB is not permanent, i.e. the treatment is ineffective and the pollutants are released more or less fully to the aquifer downstream of the barrier at some later time (curve 2), the problem has merely been shifted into the future. The ideal and desired case is when the treatment is effective and the pollutant concentration never exceeds the allowed limit (curve 3 in Fig. 12.1).

The residence time of contaminated groundwater within a PRB depends - among other parameters - on the dimension of the barrier. The reaction of elemental iron with

t'- t~ ,4...,

o

"6

1 2 ..................................... "........

.. 3allowed limit

Time Figure 12.1. Pollutant concentration in the aquifer at a location downstream of the contamination source: (1) no treatment, (2) ineffective PRB treatment with full remobilisation of the contaminant from the reactive material and (3) effective PRB treatment with low remobilisation from the barrier.

Regulatory and economic aspects 313

pollutants such as chlorinated hydrocarbons can be approximated by first-order kinetics. This means that the process can be characterised by a "half-life" which is the time needed for a 50% reduction in concentration. It follows that roughly three half-lives are needed for a 90% concentration reduction, and nearly 7 half-lives for a 99% reduction. Thus the kinetics of the remediation reactions determine the required barrier thickness. Due to these circumstances a combination of processes may be economically favourable when high inlet concentrations are combined with low target values for the remediation. Such combined approaches to remediation could include a phased remediation scheme starting with a pump-and-treat measure to "wash out" an initial contaminant load followed by a passive treatment phase by a PRB to deal with a residual pollution concentration above the remediation target level (McDonald, 2002).

In most cases laboratory experiments will be necessary to determine the reaction kinetics and the residence time needed to achieve a certain level of contaminant attenuation (see Fig. 12.2). Safe design will then require increased barrier dimensions (i.e. greater reactive zone or barrier/gate thickness), usually applied by means of a safety factor, to provide a degree of redundancy to ensure reliable barrier operation under field conditions.

Based on the residence time, the PRB design must incorporate an active treatment zone that is thick enough to achieve the desired reduction in the contaminant concentration. The minimum thickness can be roughly assessed by Equation (12.1) (Carey et al., 2002)

b = v x tR X SF (12.1)

where b is the thickness (m), v the groundwater flow velocity through the barrier (m/day), tR the residence time (days) and SF a safety factor.

The residence time can be estimated from kinetic investigations (i.e. increased-duration batch tests) and then tested in column experiments. The reactive material is usually present in a barrier in large excess so that the kinetics is of pseudo first-order (Simon et al., 2003) and can be calculated by (Carey et al., 2002)

--ln(CT/Co) tl{ = (12.2)

k

where Co is the contaminant concentration entering the PRB (mg/l), CT the target concentration down-gradient of the barrier (mg/l) and k the apparent pseudo first-order rate constant (day-l). This method will predict a rather thin barrier for very small

r

._o L r

r 0 r

0

o O

.._1

~ , . ~ fast kinetics s!ow.

" ~ ~ l n e t l c s

\ limit

Time

" i determines barrier thickness

Figure 12.2. Relationship between PRB reaction kinetics and barrier thickness.


Figure 12.3. A contaminant front moves through the barrier. If the removal rate is constant the concentration down-gradient to a barrier increases with time due to consumption of reactive material in the barrier (marked in bright grey).

flow velocities. For example, it suggests that a barrier thickness of less than 1 cm is required to reduce the uranium concentration in contaminated groundwater from Co = 3 mg/l to a target level of cT = 0.02 mg/l for a flow velocity of 0.1 m/day and k = 95 day-1 (Simon et al., 2003). However, such a calculation ignores the possibility that the flow velocity through the barrier could be higher than assumed in the calculation, e.g. due to the development of preferential flow paths in the barrier material, so that the barrier thickness becomes significantly underestimated. Also, inhomogeneous filling of the reactive material in the barrier as well as consumption or exhaustion of the material is not considered in this simplified approach.

It has been shown that where the reactive material reaches its sorption capacity or is consumed by treating the contaminants, the contaminant front moves through the reactive layer and the effective barrier thickness reduces with time (Simon et al., 2003), as illustrated in Figure 12.3. The breakthrough can be calculated from the retardation factor of the contaminant flow. The designed thickness of the barrier has to ensure that the outlet concentration is below the target concentration (Fig. 12.3).

Another technical issue is the position of the barrier relative to the contaminant source. A PRB placed immediately downstream of a pollution source will not be able to treat the contaminant plume that has already developed (Fig. 12.4). However, a PRB that is

Figure 12.4. Cleaned groundwater leaves the barrier, but an untreated zone remains upstream, and a residual contamination plume continues to migrate downstream.


positioned at a great distance downstream of the source will leave a large untreated zone between the pollution source and its treatment facility (R6ttgen, 2001). Unless the contaminant source is fresh, or the barrier is placed a long way from the pollution source, there will be a residual plume downstream of the barrier that will slowly migrate leaving a growing area of residual contamination where the contaminant concentration may be above the target value.

2. Approval of permeable reactive barriers for groundwater remediation

In most countries the installation of a PRB will be subject to a regulatory approval procedure. Since there are only a limited number of large-scale PRB installations, regulators have little experience with PRB systems. So far most PRB installations have been constructed in the USA. The governmental Environmental Protection Agency (US EPA) promotes this passive technology and even adopts the risk of failure for these systems. In such an event a facility based on active treatment processes (e.g. pump-and- treat) would be installed as an alternative (Sivavec, 2001).

In Germany, the approval of PRBs as groundwater remediation systems depends on several different laws and their respective ordinances (Steiner, 2001):

�9 building legislation; �9 water supply and management legislation; �9 soil protection legislation.

According to the German building legislation there is a right to build remediation systems both above ground and underground provided they do not put public safety at risk. The installations have to be planned and implemented according to the technical state of the art. Applications have to be filed with the local building authorities for approval. In general, building legislation should not be problematic concerning the implementation of PRB systems.

A more important issue in determining whether a PRB can be implemented in Germany is the water protection legislation. If there is interference with the direction or velocity of the groundwater flow or groundwater is extracted or re-injected, regulatory approval is usually required under the Water Supply and Management Act (Wasserhaushaltsgesetz, Federal Government, 2002). This is particularly important for funnel-and-gate systems where the water is re-directed (by the funnel walls), but the legislation will apply to most types of PRB systems as they all effect the groundwater chemistry and can potentially have adverse effects on groundwater quality. Responsible regulatory bodies are usually the local water management authorities.

The German Soil Protection Act and its accompanying ordinance also cover aspects of groundwater protection (Federal Government, 1998, 1999). These require that measures be taken if the pollutant concentration in the groundwater exceeds certain limits. In Annex 3 of the ordinance, requirements concerning remediation investigations and remediation plans are described. Under this legislation problem owners, remediation companies and the responsible authorities have to co-operate closely and are required to agree on the remediation plan. The agreement should also include the list of parameters to be monitored during operation and the monitoring intervals.


C. Economic aspects

l . P R B costs

The main decision-making tool when selecting a groundwater remediation technique is an economic evaluation of all viable techniques. During the planning stage of any kind of remediation scheme the costs and effectiveness of the different techniques have to be compared. It is very important that long-term remediation costs are factored into these comparisons, although these are difficult to estimate. In general, the costs of remedial activities are classified as capital investment costs which stem from the construction of the remediation system and operation and maintenance costs which arise continuously or periodically during operation of the system (Btirger et al., 2003). An economic evaluation of a PRB scheme needs to take account of:

�9 expenditure on the site investigation; �9 costs of preliminary and feasibility studies; �9 planning and engineering; �9 construction costs; �9 costs for reactive materials (including any recovery, replacement and disposal costs); �9 maintenance; �9 monitoring (to verify long-term performance and demonstrate remediation success).

PRB systems usually have relatively high capital investment costs (planning and construction), whereas operation and maintenance costs are normally much lower than those of most active remediation techniques, although this depends on case-specific factors such as the type of contamination, the hydraulic conditions on site, etc. (Bayer et al., 2002). From both a performance and economic standpoint the long-term performance of a PRB system is a critical area in determining its viability. Before a decision on installing a PRB system can be taken, a reliable prognosis on the probable operational life-time of the system has to be made. Indeed, lowering the expected operational and maintenance costs is the main motivation for research and development in the field of PRBs.

A number of papers have been published dealing with costs of groundwater remediation by both active (pump-and-treat) and passive (PRB) systems (Manz and Quinn, 1997; Teutsch et al., 1997; Reeter et al., 1999; Stupp, 2000; Edel and Voigt, 2001; FRTR, 2001; Bayer et al., 2002; BUrger et al., 2003). However, the projects differ substantially in size and complexity. In a report of the US EPA (Powell et al., 2002) the economics of 22 PRB systems were analysed. The volume of groundwater treated was between 1OOO and some 400,000 m 3 per year. The operating costs and capital costs in US$ per m 3 water treated per year were highest for the smaller remediation projects and decreased significantly for large projects. Of course these costs also depend on the site characteristics and the type of contaminant. Table 12.1 lists data from the report, and Figure 12.5 displays the data as a function of the groundwater volume treated per year in the respective plants (this comparison makes no allowance for the different design lives of the different installations).

Remediation costs for chlorinated hydrocarbons in groundwater using PRBs were compared with the pump-and-treat method in a report of the US Department of Defence (Reeter et al., 1999). The cost saving for the example discussed in this report was


Table 12.1. Cost data from PRB projects in Northern America (based on data given in Powell et al., 2002): annualised capital costs per m 3 of treated groundwater (GW), normalised to 1 year of operation, and annual operation and maintenance costs (O&M).

Site GW volume Annualised PRB O&M Annual PRB (m3/year) capital (S/year) O&M ($/m 3)

cost ($/m 3)

USCG Support Center 15,064 52.77 85,000 5.64 Intersil Site 6624 134.67 95,000 14.34 Watervliet Arsenal 2945 125.65 Moffett Federal Airfield 4504 134.85 90,000 19.98 Somersworth Landfill SF Site 412,565 5.91 Dover AFB, DE 4921 134.34 Kansas City Plant, MO 23,770 61.00 Aircraft Maintenance, OR 58,289 18.61 50,000 0.86 Caldwell Trucking, NJ 102,195 Former Manufacturing, Fairfield, NJ 19,682 44.46 25,000 1.27 Industrial Site, Coffeyville, KS 189,629 2.11 Industrial Site, NY 41,257 24.24 Industrial Site, SC 13,626 29.36 Nickel Rim, Ontario 939 95.88 30,000 31.96 Cape Canaveral, FL 2324 327.09 MMR CS- 10 Plume, MA 49,584 Pease AFB, NH 35,000 Warren AFB Spill Site 7, WY 79,107 27.81 London, Ontario 3000 Moffett, full-scale est. 179,031 27.17 72,278 0.40 Dover, AFB full-scale est. 9879 89.59 148,000 14.98

Note: for some projects not all the required data are available.

350

300

r E 250

200 0 ~ 150 e~ ~ 100

o~ ~

* Capital costs o O&M costs

50 ~** �9 o � 9 �9

0 i i I 0.1 1 10 100

Treated GW volume (1000 m3/a)

35

30

25,~ E

20

O 15 o

o~ 10 o

5

0 1000

Figure 12.5. PRB capital investment costs (US$ per m 3, normalised to 1 year of operation) and annual operation and maintenance (O&M) costs per m 3 groundwater as a function of the treated volume per year (Powell et al., 2002).


between 50 and 70% over the long term, as can be seen in Figure 12.6. In this comparison barrier maintenance costs of 268,000 US$ every 10 years were adopted. This was equivalent to 25% of the iron costs and was actually only a best guess because no long-term experience with PRBs then existed. It is thus uncertain whether or not such cost savings can indeed be achieved.

2. Reactive materials

The performance of a PRB is dependent on the effectiveness of the reactive materials used. The main criteria upon which reactive materials are selected are:

�9 reactivity; �9 stability; �9 availability; �9 costs; �9 hydraulic performance; �9 environmental compatibility; �9 safety.

The ideal reactive material has a constant high permeability, high reaction rate, long- term stability and low acquisition costs, and is non-hazardous to both construction personnel and the environment. When the reactive material used in a PRB meets these criteria, the barrier will almost certainly receive a positive economic evaluation.

A number of reactive materials have been investigated for their effectiveness at removing uranium (and some other heavy metals such as arsenic) from contaminated groundwater. Of these materials, elemental iron, hydroxyapatite, activated carbon and a new material, PANSIL, were most effective in uranium retention. Average costs for a variety of reactive materials are listed in Table 12.2.

An anion-exchange resin (Varion AP) has also been tested with uranium-contaminated groundwater, and is also included in Table 12.2. It has been shown that this exchanger, with functionality based on 2,N-dimethyl-pyridinium groups, has an excellent sorption capacity for anionic uranyl carbonate species. As this anion exchange resin has been

40000 35000

~'30000 25000

~20000 15000

O o 10000

5000 0

S J

f

I I i I I I ! l i

0 10 20 30 40 50 Years

I--o-PRB -me- Pump & Treat I

Figure 12.6. Cost comparison of groundwater remediation by pump-and-treat technology and PRBs (Reeter et al., 1999).


Table 12.2. Average costs of reactive materials tested applicable for passive groundwater remediation.

Reactive material Costs

Activated carbon Elemental iron Zeolites Hydrated lime Limestone Hydroxyapatite PANSIL b Varion AP Ferrosorp

1280 EUR/ton (1036 US$/ton) a 350 US$/ton (EnviroMetal Technologies Inc.) 450 US$/ton (Bowman et al., 1999) 100 US$/ton (Murphy et al., 1999) 68 US$/ton (Ott, 2000) 500 US$/ton (Bostick et al., 2000) 660 US$/ton (Bryant et al., 2003) --~ 3300 EUR/m 3 (2670 US$/ton) a 1500 EUR/ton (Kiel3ig, 2000) (1214 US$/ton) a

a Calculated by 0.8 EUR/US$. b For details on PANSIL see Chapter 7.

widely used in uranium processing for removing uranium from mine water, it can be readily recommended for PRB applications. The cost of the resin is high, but this cost does not consider that the resin can be re-used after being regenerated (additional investigations are needed to determine the best regeneration technology for exhausted resins). An advantage of using ion exchange resins is that they do not have much effect on the general water chemistry, but they do have the disadvantage that their use is only likely to be economical within a reactor-type installation.

3. Economic aspects of the experimental barrier installation in P~cs, Hungary

The investment costs for the pilot installation in Prcs are presented in Table 12.3. The total cost is approximately 36,300 EUR. This sum includes all the main costs for the installation (e.g. acquiring the reactive material, excavating the trench and building the monitoring network), but not the cost of the site investigation and earthworks to prepare the ground. This PRB was constructed as an experimental test facility, therefore the operation and maintenance costs have not been estimated. Material costs include the cost for 38 t of elemental iron used in the barrier.

The specific investment cost is approximately 50 EUR/m 3 (see Table 12.3), and approximately 41% of that sum was spent on materials. These cost figures appear to be

Table 12.3. Investment cost of the Prcs experimental barrier.

Groundwater flow Investment costs rate (m3/a) (EUR)

Material Construction Total

Specific inv. costs (EUR/m 3 groundwater)

750 15,000 21,300 36,300 48.40

Dimensions of PRB: length, 6.9 m; depth, 3.8 m; thickness, 1.5 m.


within the normal range for PRBs despite the barrier length being less than 7 m, which is small compared with most barrier installations (Bayer et al., 2002). The total costs for the installation would have been higher if the barrier had been built in a country with higher unit labour costs, or if pure elemental iron had been used in the barrier instead of an iron/ sand mixture.

D. Outlook

PRB technology is a promising approach for the integrated management of polluted groundwater. Therefore a frequently asked question is whether a PRB is the solution to a particular groundwater contamination problem. To answer this question, it is important to reconsider the following points:

�9 Basic concept o f PRBs. Is a passive remediation approach suitable for solving the pollution problem within the given remediation objectives for a particular site (e.g. time frame, future land-use, regulatory provisions, etc.)?

�9 Potential applications. Is the use of the PRB technique feasible in terms of construction and site management?

�9 Characteristics o f the contaminated site. Are the geological conditions, and the extent and concentration of the contamination suitable for implementation of a PRB system?

�9 Available methods. For a given contaminant and groundwater composition, are there contaminant removal processes (e.g. reductive mechanisms, sorption) and suitable reactive materials available for use in the planned PRB?

These are similar questions and considerations to those associated with all other remediation options. Indeed, PRB technology is not appropriate for all cases of groundwater contamination, just as other methods are not universally applicable. The decision on whether to implement a PRB system or another remediation method is case- specific and depends mainly on the comparison of feasibility and costs.

References

Bayer, P., Btirger, C., Finkel, M., Teutsch, G., 2002. Technical and economical comparison between funnel-and-gate and pump-and-treat systems: an example for contaminant removal through sorption. In: Simon, F.G., Meggyes, T., McDonald, C. (Eds), Advanced Groundwater Remediation - Active and Passive Technologies. Thomas Telford, London, pp. 267-282.

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.

Bowman, R.S., Zhaohui, L., Roy, S.J., Burt, T., Johnson, T.L., Johnson, R.L. 1999. Surface-altered zeolites as permeable barriers for in-situ treatment of contaminated groundwater. Department of Earth and Environmental Science, New Mexico Institute of Mining and Technology, Phase II Topical Report, DE-AR21-95MC32108.

Bryant, D.E., Stewart, D.I., Kee, T.P., Barton, C.S., 2003. Development of a functionalized polymer- coated silica for the removal of uranium from groundwater. Environ. Sci. Technol. 37, 4011-4016.

Btirger, C., Finkel, M., Teutsch, G., 2003. Reaktionswandsysteme und "pump-and-treat" - ein Kostenvergleich. Grundwasser 3/2003, 169-180.


Carey, M.A., Fretwell, B.A., Mosely, N.G., Smith, J.W.N., 2002. Guidance on the use of permeable reactive barriers for remediating contaminated groundwater. Environment Agency, National Groundwater & Contaminated Land Centre Report, NC/01/51, Solihull, UK, 140 pp.

Edel, H.G., Voigt, T., 2001. Aktive und passive grundwassersanierung - ein Verfahrens- und Kostenvergleich. TerraTech 1/2001, 40-44.

EnviroMetal Technologies Inc., http://www.eti.ca/. Federal Government, 1998. Federal Soil Protection Act (Bundes-Bodenschutzgesetz - BBodSchG),

March 17, 1998, amended September 9, 2001. Bundesgesetzblatt, I, Bonn, Germany, p. 2331. Federal Government, 1999. Federal Soil Protection and Contaminated Sites Ordinance (Bundes-

Bodenschutz- und Altlastenverordnung - BBodSchV), July 12, 1999. Bundesgesetzblatt, 1, Bonn, Germany, p. 1554.

Federal Government, 2002. Water Supply and Management Act (Wasserhaushaltsgesetz - WHG), August 19, 2002. Bundesgesetzblatt, I, Bonn, Germany, p. 3245.

Federal Remediation Technologies Roundtable (FRTR), 2001. Cost analysis for selected groundwater cleanup projects: pump and treat systems and permeable reactive barriers. US Environmental Protection Agency, EPA 542-R-00-013.

Kie[3ig, G., 2000. Entwicklung von Verfahrenstechnischen LiSsungen zur Sanierung von Grundwas- serscNiden im Abstrom von Absetzbecken der Uranerzaufbereitung Mittels Permeabler Reaktiver W~inde. Wismut GmbH, Abschlu[3bericht, 02WB9891/8 BMBF.

Manz, C., Quinn, K., 1997. Permeable treatment wall design and cost analysis. Proceedings of the International Containment Technology Conference, February 1997, St Petersburg, FL, pp. 788-794.

McDonald, C., 2002. Discussion: status, directions and R&D issues. In: Simon, F.G., Meggyes, T., McDonald, C. (Eds), Advanced Groundwater Remediation - Active and Passive Technologies. Thomas Telford, London, pp. 303-325.

Murphy, N.C., Taylor, J.R., Leake, M.J., 1999. Coming to terms with acid drainage. State of the Art in Mining Environmental Management, Manila, Philippines.

Ott, N., 2000. Permeable reactive barriers for inorganics. US Environmental Protection Agency, Internet Report, tdott_prb.pdf on http://www.clu-in.org.

Powell, R.M., Powell, P.D., Puls, P.D., 2002. Economic analysis of the implementation of permeable reactive barriers for remediation of contaminated ground water. US EPA, National Risk Management Research Laboratory, Project Report, EPA/600/R-02/034, Cincinnati, OH.

Reeter, C., Chao, S., Gavaskar, A., 1999. Permeable reactive wall remediation of chlorinated hydrocarbons in groundwater. US Department of Defense, Environmental Security Technology Certification Program (ESTCP), ESTCP Cost and Performance Report.

R6ttgen, K.P., 2001. Akzeptanz von Reinigungsw~inden aus fachbeh6rdlicher Sicht. Reinigungsw~inde auf dem Vormarsch, Magdeburg.

Simon, F.G., Segebade, C., Hedrich, M., 2003. Behaviour of uranium in iron-bearing permeable reactive barriers. Sci. Total Environ. 307, 231-238.

Sivavec, T., 2001. GE Corporate Research and Development, personal communication. Steiner, N., 2001). Rechtliche Grundlagen und genehmigungsrechtliche Aspekte. Reinigungsw~inde auf

dem Vormarsch, Magdeburg. Stupp, H.D., 2000. Grundwassersanierung von LCKW-Sch~iden durch pump and treat oder reaktive

systeme. TerraTech 9 (2), 34-38. Teutsch, G., Tolksdorff, J., Schad, H., 1997. The design of in situ reactive wall systems - a combined

hydraulical-geochemical-economical simulation study. Proceedings of International Containment Technology Conference, February 1997, St Petersburg, FL, pp. 917- 924.

Subject index

abiotic reductive degradation, 5 activated carbon, 2, 7-9, 18-9, 67, 184,

305-309 adsorption, 3, 7, 19, 291,294, 306 adsorptive reactor, 283 alkalinity, 6, 12- 5, 17 amalgamation, 6 Amberlite IRN77, 172, 179t, 180t anion exchange capacity, 7 anode, 184 antimony, 8 apatite minerals, 79 autunite, 64, 79-80

barrier thickness, 5, 314 batch exposure tests:

Amberlite IRN77 (cation exchange resin), 176

Methodology, 161 PANSIL, 168, 169 Varion AP (anion exchange resin), 175

bentonite suspension, 28 bicarbonate, 10, 15, 138, 150 biobarriers, 45 biodegradation, 46 biofilm, 307 biofouling, 307 biomass, 11, 16 biopolymer trenching, 48 Bitterfeld, 18 Borden, 14, 18 bored-pile cut-off walls, 32 brownfield, 283 BTEX, 284, 305, 306f bulk precipitation process, 65, 71

calcite, 12, 303, 304, 304f capital costs, 317t carbonate, 6-7, 10-15, 150, 304, 304f

cathode, 183 cation exchange capacity, 7 chemical resistance, 34 chernikovite, 79-80 chlorinated hydrocarbons, 5, 14-5, 284, 292 chloroethene, 5 chromate, 2, 6, 8 clay lenses, 186, 195 clay, 7-8, 195, 227, 229, 272 clay-cement-water mix, 31 clinoptilolite, 55, 71 clogging, 184, 200, 207, 294, 303-4, 307 coating, 9-10, 16, 18, 184, 200, 207, 279, 281,

307 column experiment, 4-5, 11, 13, 77, 82-4, 84t,

85-7, 87t, 88t, 94-6, 98, 104, 111, 112f, 113t, 114, 114t, 115, 115f, 116t, 127-8, 137, 160, 163t, 261,313

column exposure tests: Amberlite IRN77 (cation exchange resin),

175 Methodology, 162- 3 PANSIL, 167, 171 Varion AP (anion exchange resin), 177

combined PRB - phytoremediation systems, 47 composite cut-off wall, 30 compost, 6, 17-8 compressive failure strain, 34 conceptual design report, 36 conglomerate, 225 containment, 4, 286, 290 continuous barriers, 3f, 4, 12, 14, 16, 18,

137, 139, 261 continuous reactive barrier, 3, 137 copper, 6 corrosion, 5, 10-6, 143 cost analysis, 48 costs, 206 cut-off slurries for single-phase walls, 33

324 Subject index

cut-off wall chamber system, 30 cut-off wall materials, 32 cut-off wall technology for PRBs, 38

Darcy' s law, 118 deep aquifer remediation tool (DART), 4,

27, 42 deep soil mixing, 40 degradation, 3, 5, 13, 15-6, 19, 291,301,304 dehalogenation, 5, 10, 12-4 denitrification, 17-8 design considerations, 35 diatomite, 7-8 dichloroethene, 5 diffusion, 184, 188, 191 - 3, 201 dissolved iron, 139-40, 150-1 distribution coefficient, 98 distribution coefficient, Rd, 179

economic evaluation, 316 electric conductivity, 144, 268f, 275f electrodes, 183, 186, 188, 205, 229, 229f electrokinetic cell, 186-7, 187f electrokinetic fence, 184, 201,203, 205-7 electrokinetic soil remediation, 183 electrokinetic techniques, 183 electrokinetics, 19, 183 electrolysis, 183, 191-2 electromigration, 183, 185, 193, 203 electroosmosis, 183, 185, 195 electrophoresis, 183 elemental copper, 6, 96 elemental iron, 2, 5, 9-11, 68, 137-8, 184, 189,

201-2, 262, 267 environmental compatibility, 8 extra-cellular polymer (EPS), 45

feasibility study, 4, 291-2, 308 floor-scale tests, 116 flow rate, 137-9, 143, 151 frozen walls, 32 Fry Canyon, 77 full-scale, 9, 11, 12-6, 19, 283, 307 funnel-and-gate, 2-3, 3f, 11-2, 14-5, 137,

139, 228

geochemical modelling, 99, 121, 154, 163-4, 165f, 169-70

geochemical simulation, 131 geoelectrical methods, 229 geological structure, 226, 229 geomembrane, 262- 3 geophysical investigation, 229 Geosyphon, 4 geosynthetic clay liners, 262-3 gypsum, 17

heap leaching, 211,214f, 218f horizontal reactive barrier, 217 hydrated lime, 67, 71 hydraulic barrier, 283-4, 288f, 292, 295 hydraulic conductivity, 8, 16-7, 150,

235, 294, 302 hydraulic fracturing, 44 hydraulic properties, 127 hydrogen gas, 143 hydroxyapatite (HAP), 8, 59, 79, 80f, 82, 137,

140, 143, 150

immobilisation, 2-3, 6-7 industrial area, 4 injected systems, 44 injection walls, 32 inoculum, 45 ion exchange, 3, 6-7, 80, 219 ion exchange resins, 153, 172-3,

179-80 iron, 5, 9-10, 16, 137-8, 261-2, 263f iron barrier, 9, 12, 15-6, 150 iron corrosion, 69 iron oxidation, 9, 79 iron oxides, 70-1 iron sponge, 129f

jet cutting head, 41 jet grouting, 32 jet pump, 41 Jet technology, 41

Killarney, 18 kinetics, 92

GAC, 9, 18-9 gate structures, 39 gates, 3-4, 11, 14-5

"Lasagna" process, 184 lead, 6, 8 legislation, 315

Subject index 325

lifetime, 10, 13, 15-6, 19, 261,294 linoleum production, 283-4, 286 loess loam, 189, 190t, 195 long-term behaviour, 9, 184, 279, 283, 303 long-term performance, 9-10, 13-4, 16-9,

207, 301

mackinawite, 12, 17 mercury, 6 microbial degradation, 291,304 mine effluent, 1 mine water, 2, 5-6, 211,216 monitoring, 212, 218, 221,267, 280, 283, 292,

294-6, 299, 300f, 30If, 308 monitoring wells, 137-8, 222, 233, 264, 265f,

266f, 267, 270-1,281 Monticello, 13

natural attenuation, 1, 19, 301 Nickel Rim Mine, 16 nitrate, 6, 10, 12, 14, 16-7, 18

operation and maintenance costs (O&M), 317t oxygen, 10, 15, 18, 143, 150, 301,307-8

PAH, 284, 286, 287t, 292, 307-8 PANSIL,

chemical structure, 155, 160 efficiency and durability, 160 mechanism of contaminant removal, 157,

160 synthesis, 155

passive treatment, 1 P~cs, 82, 111-3, 116, 118, 130, 137-8, 150,

160, 163, 163t, 164, 165f, 170, 171f, 172, 177, 190, 211, 261, 263f, 264f, 270, 280, 319, 319t

P~cs groundwater, 163, 163t perchloroethene (PCE), 5, 13, 15 permeability, 7-8, 10-3, 118, 143, 150, 236t,

288, 294, 301-3, 307-8 pH, 137, 139, 143f, 215, 267, 268f, 271f,

275f phenol, 284, 287t, 292, 294 pillared clay, 8 pilot-scale, 9, 11, 15, 17, 19, 137, 215, 242 polycyclic aromatic hydrocarbons, 284 polymer films, 45

porosity, 7, 10-7, 185, 276, 281,301,304, 306f

porosity loss, 11-6, 200, 204 PRB costs, 316 precipitates, 9-17, 150, 190, 201f, 270,

276, 281 precipitation, 3, 6-8, 10-7, 19, 139, 150,

199-200, 204-5, 269-70, 276, 281, 303-4, 304f, 306-7

preferential flow, 302-4, 306 preferential pathways, 261 pump-and-treat, 1, 9, 217, 220, 229, 290 pyrite, 222

quartz sand, 189, 189t

radiotracers, 86 rate constant, 93, 97 reaction kinetics, 6-7, 15 reactive components, 2 reactive material, 2-3, 5, 7, 137, 150, 294, 301,

303, 306 reactive thin walls, 39 reactive zone thickness, 4 redox conditions, 138 redox potential, 301 redox reactions, 2 reductive barrier, 3 residence time, 4-5, 8, 13, 16-8, 138-9, 143,

143f, 150, 302, 313 retardation, 7, 98 retention, 7, 196, 198, 200 Rheine, 13 risk management, 1 RUBIN, 9

safety factor, 5 sandstone, 224, 238 saturation Indices, 123 sawdust, 6, 18 scanning electron microscopy (SEM), 120f, 304 schoepite, 67, 70 secondary minerals, 6, 150, 183, 207 self-hardening slurry, 28 sheet-pile boxes, 39 sheet-pile walls, 31 siderite, 10- 2, 17 single-phase diaphragm wall, 28 siphon, 138

326 Subject index

slurry jets, 41 Soil Protection Act, 315 soilcrete columns, 32 solubility products, 99 sorption, 2-3, 7-8, 18-9, 291,306 sorption barrier, 3, 7-8, 18 sorption isotherms, 7

Amberlite IRN77 (cation exchange resin), 174

Methodology, 161 PANSIL, 166, 166f Varion AP (anion exchange resin), 174

stability, 8 steel fibres, 138, 140, 144t, 150 stop-end tubes, 28 stratigraphy, 238 subsurface biofilm barriers, 45 sulphate reduction, 2, 6, 10, 16-8 sulphate, 150 sulphide, 6-7, 10-2, 15, 17, 150 Sunnyvale, 14 surface charge, 57, 66, 71 surface sorption, 80 surface speciation diagram, 62, 71

tailings, 211,217, 219, 220f, 221f, 226, 227f tar factory, 283-4 tetrachloroethene (TCE), 5, 12- 3, 16, 19,

287t thin wall slurries, 33 thin walls, 31 total dissolved solids (TDS), 144, 150, 215,

220f, 268f, 269, 271,272f, 273f, 276, 281

tracer, 13-4, 301-2, 303f, 303t transmissivity, 236t trichloroethene, 5, 184 two-phase cut-off slurries, 33 two-phase diaphragm walls, 28

ultrasonic, 16 ultrasound, 16 uniaxial compressive strength, 34 uraninite, 70, 78 uranium phosphates, 79 uranium, 6, 8, 12-3, 43, 53-6, 56f, 57, 57f,

58-9, 59f, 60, 60f, 61f, 63, 63f, 64, 64f, 65-6, 66f, 67, 67f, 68, 68t, 69, 69f, 70-2, 72f, 73, 76-81, 81f, 82-6, 86f, 87, 87t, 88, 88t, 89-90, 90t, 91-3, 93f, 94, 94f, 95-6, 96f, 97-100, 111-3, 113t, 114, 114t, 115, 115f, 116, 116t, 119-20, 121f, 122f, 123, 123f, 124, 124f, 125, 125f, 126, 126f, 127, 127f, 129-32, 137, 153-5, 157-9, 159f, 160-3, 168f, 170, 173-5, 175f, 176, 176f, 177-8, 180-1,211,261, 262f, 267, 268f, 270-1,272f, 275f, 281, 314,318,319

uranyl, 2 uranyl measurement, 160 urban areas, 4, 9, 283

vinylchloride, 5 Varion AP, 172

waste rock, 211, 215 Water Supply and Management Act, 315 well-based systems, 42 wetland system, 1 wood chips, 6, 17

X-ray diffraction, 304

yellow cake, 211

zeolite, 6-9, 55 zero-valent iron (ZVI), 5, 137 zeta potential, 57f

long-term performance of permeable reactive barriers, volume 7 (trace metals and other contaminants...

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