COMMITTEE ON EPA 542-R-98-003

THE CHALLENGES OF May 1998

MODERN SOCIETY www.clu-in.com

www.nato/ccms

NATO/CCMS Pilot Study

Evaluation of Demonstrated and

Emerging Technologies for the

Treatment of Contaminated Land

and Groundwater (Phase III)

1998

SPECIAL SESSION

Treatment Walls and

Permeable Reactive Barriers

Number 229

NORTH ATLANTIC TREATY ORGANIZATION
NATO/CCMS Pilot Study

Evaluation of Demonstrated and

Emerging Technologies for

the Treatment of Contaminated

Land and Groundwater — Phase III

SPECIAL SESSION ON

Treatment Walls and

Rermeable Reactive Barriers

Harald Burmeier, Chairman

University of Vienna

Vienna, Austria

FEBRUARY 22 - 28, 1998

May 1998
NOTICE

This report was prepared under the auspices of the North Atlantic Treaty Organization’s Committee on the

Challenges of Modern Society (NATO/CCMS) as a service to the technical community by the United States

Environmental Protection Agency (U.S. EPA). The document was funded by U.S. EPA’s Technology

Innovation Office under the direction of Dr. Michael Kosakowski. The Annual Report was edited and

produced by Environmental Management Support, Inc., of Silver Spring, Maryland, under U.S. EPA

contract 68-W6-0014. Mention of trade names or specific applications does not imply endorsement or

acceptance by U.S. EPA.
Treatment Walls NATO/CCMS Pilot Study Phase III

i

CONTENTS

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iii

General Overview .................................................................. 1

Harald Burmeier

Permeable Reactive Barrier Research at the National Risk Management Research Laboratory,

U.S. Environmental Protection Agency ...........................................3

Robert W. Puls

Technical Construction of Treatment Walls

Permeable Treatment Walls—Design, Construction, and Cost ................................6

Eberhard Beitinger

Development of Iron-Based Reactive Barrier Technologies for Remediation of Chlorinated Organic

Contaminants in Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

Robert W. Gillham

Practical Solutions for the Treatment of Polluted Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22

Gérard Evers

Reactive Materials

Degradation of TCE at Zero-Valent Iron: Chemical Processes Effecting the Design and Performance

of Permeable, Reactive Fe(0) Walls . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30

Wolfgang Wüst, O. Schlicker, and A. Dahmke

The Treatment of Groundwater with Mixed-Wastes: Reductive Dechlorination of 

TCE and Reductive Precipitation of Uranium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

Liyuan Liang and B. Gu

Bioprocesses in Treatment Walls: Bioscreens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44

Huub H. M. Rijnaarts

Novel Catalyses for Reactive Barriers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48

Timothy M. Vogel

Full-Scale Projects

Funnel-and-Gate Systems for In Situ Treatment of Contaminated Groundwater at 

Former Manufactured Gas Plant Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

Hermann Schad and Peter Grathwohl

Reactive Treatment Zones: Concepts and a Case History . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 66

Stephan A. Jefferis and Graham H. Norris
Treatment Walls NATO/CCMS Pilot Study Phase III

ii

Horizontal Treatment Barriers of Fracture-Emplaced Iron and Permanganate Particles . . . . . . . . . . . . 77

Robert L. Siegrist, Kathryn S. Lowe, Lawrence W. Murdoch,

Traci L. Case, Douglas A. Pickering, and Thomas C. Houk

In Situ Remediation Research in a Complexly Contaminated Aquifer: The SAFIRA Test Site 

at Bitterfeld, Germany . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84

H. Weiss, F.-D. Kopinke, P. Popp, and L. Wünsche

Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 92

Harald Burmeier

ABOUT THE AUTHORS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 95

NATIONAL CONTACTS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 98

PARTICIPANTS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102
Treatment Walls NATO/CCMS Pilot Study Phase III

iii

Introduction

The Council of the North Atlantic Treaty Organization (NATO) established the Committee on the

Challenges of Modern Society (CCMS) in 1969. CCMS was charged with developing meaningful programs

to share information among countries on environmental and societal issues that complement other

international endeavors and to provide leadership in solving specific problems of the human environment.

A fundamental precept of CCMS involves the transfer of technological and scientific solutions among

nations with similar environmental challenges.

The management of contaminated land and groundwater is a universal problem among industrialized

countries, requiring the use of existing, emerging, innovative, and cost-effective technologies. This

document provides a report from the first meeting of the Phase III Pilot Study and is designed to share

information among countries on innovative treatment technologies. The United States is the lead country

for the Pilot Study, and Germany and The Netherlands are the Co-Pilot countries. The first phase

successfully concluded in 1991, and the results were published in three volumes. The second phase, which

expanded to include newly emerging technologies, concluded in 1997; final reports documenting 52

completed projects and the participation of 14 countries will be published in 1998. Through these pilot

studies, critical technical information has been made available to participating countries and the world

community. 

Phase III focuses on the technical approaches for addressing the treatment of contaminated land and

groundwater. This includes issues of sustainability, environmental merit, and cost-effectiveness in addition

to continuing to draw on the merits of emerging remediation technologies. The objectives of the study are

to critically evaluate technologies, promote the appropriate use of technologies, use information technology

systems to disseminate the products, and to foster innovative thinking in the area of contaminated land. The

first meeting of the Phase III Pilot Study on the Evaluation of Demonstrated and Emerging Technologies

for the Treatment and Clean Up of Contaminated Land and Groundwater convened in Vienna, Austria, on

February 22-27, 1998, with representatives of 20 countries attending. Each participating country presents

case studies or projects to the Pilot Study for discussion. 

Also at this first Phase III meeting, the first special technical session was convened on treatment walls and

related permeable reactive barrier technologies, under the chairmanship of Prof. Dipl.-Ing. Harald Burmeier

of the University of Northeast Lower Saxony. This report contains the proceedings of that special session.

A companion publication, the first Annual Report of the Phase III studies, contains abstracts of the first 15

demonstration projects selected, reports on the legislative, regulatory, programmatic, and research issues

related to contaminated land in each participating country, and the statement of purpose for the Phase III

Pilot Study. General information on the NATO/CCMS Pilot Study may be obtained from the Country

Representatives listed at the end of this report, or—for each paper—from the authors themselves.

Stephen C. James

Walter W. Kovalick, Jr., Ph.D.

Co-Directors
Treatment Walls NATO/CCMS Pilot Study Phase III

 University of Applied Studies and Research, University of Northeast Lower Saxony, Herbert-Meyerstrasse 7, 29556 1

Suderburg, Germany, tel: 49/5103-2000, fax: 49/5103-7863, e-mail: h.burmeier@t-online.de

1

General Overview

Harald Burmeier1

Remediation of groundwater is most commonly performed by “pump-and-treat” methods: the contaminated

groundwater is extracted from the aquifer and subjected to above-ground-treatment before being reinjected

or discharged. Multi-component systems are designed according to the particular contamination by taking

physical, chemical, and biological treatment processes into consideration.

Pump-and-treat methods require continuous energy for pumping water from the extraction wells and

operating the water treatment systems. Besides that, periodic maintenance and monitoring has to be

performed. Generally, these systems have to be operated for a long time. As a consequence, remediation

costs are substantial. Despite the wide spectrum of technologies available for above-ground treatment,

residual contaminants frequently remain at undesirably high levels in the subsurface. 

Alternatives to these methods are passive working systems that depend on the usage of the groundwater’s

natural hydraulics. There is a wide spectrum of possible solutions ranging from intrinsic remediation to

permeable reactive or adsorptive barriers. 

As early as 1982, the idea to treat contaminated groundwater by installing a permeable reactive barrier was

shown in a figure in a remediation handbook issued by the U.S. EPA. But the 1980s were not the time for

further development of such innovative in-situ clean-up methods. 

In 1989, the potential of Permeable Reactive Barriers (PRB) was recognized more clearly and was developed

further by the University of Waterloo, Canada. Since that time a lot of laboratory- and bench-scale investigations led to the first full-scale in situ demonstration at Borden, Ontario, Canada.

Now, in 1998, over 500 project studies world-wide reflect the interest in this technology. The small number

of so far only 20 commercial applications shows that it is a long way from the pilot study to the field

application. 

Today, a lot of experts in North America and Europe are working on several research issues concerning the

PRB technology, as there are:

 • processes in the media and aquifers (hydraulics, geochemistry, etc.),

 • suitable materials for reactive walls,

 • design and construction,

 • long-term effectiveness,

 • treatment of contaminant-mixtures, and

 • standards for construction, monitoring, etc.

Although many questions remain to be answered, a significant cost reduction potential of more than 30

percent is expected when PRBs are used instead of pump-and-treat methods.

In this Special Session of the Phase III Pilot Study on the Evaluation of Demonstrated and Emerging Technologies for the Treatment and Clean Up of Contaminated Land and Groundwater (Phase III), a dozen
Treatment Walls NATO/CCMS Pilot Study Phase III

2

papers are presented on the international state-of-the-art concerning technical construction issues, reactor

materials, and full-scale projects. This report summarizes those proceedings and the discussions following

them, research and development needs, and recommendations for further action items.

I would like to extend my special thanks to all experts for preparing reports and giving presentations during

this Pilot Study Meeting in Vienna. Last-but-not-least, I would like to gratefully acknowledge the input and

substantial support provided by Steve James, John Moerlins, and Volker Franzius during the preparation

and coordination of this session. 
Treatment Walls NATO/CCMS Pilot Study Phase III

 U.S. Environmental Protection Agency, R.S. Kerr Environmental Center, P.O. Box 1198, Ada, OK 74821-1198, USA, 1

e-mail: puls.robert@epamail.epa.gov

3

Permeable Reactive Barrier Research at the

National Risk Management Research Laboratory,

U.S. Environmental Protection Agency

Robert W. Puls1

Much of the current research on ground-water remediation has focused on the removal of contaminated

water from the subsurface and treating it at the surface. While the removal of the contaminants is desirable,

the costs are often prohibitive and rarely are contaminant concentrations lowered to the required regulatory

levels. This has been particularly evident for standard “pump-and-treat” approaches. In situ chemically

reactive permeable walls are being considered as a low-cost and effective alternative for the treatment of

contaminated waste sites. The chemical form of the contaminant in question is transformed via oxidation,

reduction or precipitation reactions to an immobilized or non-toxic form. The application of in situ

approaches to subsurface remediation will increase the emphasis on adequate site characterization and

thorough understanding of the subsurface system targeted for remediation as well as the geochemical

mechanisms controlling contaminant transformations.

Background

Research into the use of zero-valent metals to remediate ground-water contaminated with mixed wastes

(inorganic and organic) has been ongoing at the U.S. Environmental Protection Agency’s National Risk

Management Research Laboratory (NRMRL) in Ada, Oklahoma since 1991. The primary emphasis has been

on inorganic constituents such as chromate, arsenic, nitrate and sulfate and chlorinated organic compounds

(e.g., trichloroethylene, cis-dichloroethylene, and vinyl chloride). Laboratory research conducted by

scientists at the NRMRL conclusively demonstrated the effectiveness of using zero-valent metals to

remediate chromate and chlorinated organic compounds in groundwater and this research was scaled up to

a pilot-scale demonstration in September 1994, near an old chrome plating facility on the U.S. Coast Guard

(USCG) Support Center near Elizabeth City, North Carolina. Results indicated that complete treatment of

chromium and the chlorinated organic compounds in the groundwater was possible using this technology.

Chromium concentrations were reduced to less than 0.01 mg/l, much less than drinking water limits. The

chromate was reduced via corrosion of the elemental iron to the nontoxic and insoluble chromic ion (Cr ), 3+

which forms an insoluble mixed chromium-iron hydroxide phase. The chlorinated organic compounds were

similarly degraded to below maximum contaminant levels set for groundwater by the U.S. EPA. 

Full-Scale Systems

A full-scale demonstration of a permeable reactive barrier to remediate groundwater contaminated with

chromate and chlorinated organic compounds was initiated at the USCG site by researchers from NRMRL

and the University of Waterloo in 1995. A continuous wall composed of 100% zero-valent iron was installed

in June, 1996, using a trencher that was capable of installing the granular iron to a depth of 8 m. The

trenched wall was approximately 0.6 m thick and about 60 m long. The wall begins about 3 feet below the

ground surface and consists of about 450 tons of granular iron. The installation of the wall was completed

in less than a day. The total installation cost was $500,000, with the cost of iron representing approximately

35% of the total. Total installation, site investigation, design feasibility tests, general contracting, and postinstallation monitoring costs have been less than $1 million. Savings compared to traditional pump-and-treat
Treatment Walls NATO/CCMS Pilot Study Phase III

4

exceed $5 million over a 10-year period primarily due to reduced operation and maintenance costs due to

the passive nature of this technology.

The continuous trenching equipment used for the installation was similar to a large “Ditch Witch.” It uses

a large cutting chain excavator system combined with a trench box and loading hopper. As the trenching

machine moved along, the cutting chain excavation removed the native soil, while a front end loader filled

the hopper of the trench box. The granular iron flowed through the hopper and out the back of the trench

box into the just-excavated trench. The walls of the trench box extended to the width and depth of the trench

maintaining the opening until the iron was emplaced.

The system was designed to meet lower concentrations of contaminants below maximum concentration

limits (MCLs) for groundwater set by U.S. EPA of 0.05 mg/l Cr(VI), 5 μg/l TCE, 70 μg/l DCE and 2 μg/l

VC.

Results

To-date, there has been two years of post-installation performance monitoring performed by these same

researchers. For all but one quarterly sampling event, 15 multilevel samplers (7 to 11 sample ports per

sampler) and 9 to 10 compliance (5-cm PVC) wells have been sampled. In addition to on-site sampling of

the full suite of geochemical indicator parameters listed in the site work plan, samples have been collected

for laboratory analysis of the following constituents: TCE, cis-DCE, vinyl chloride, ethane, ethene, methane,

major anions, and metals. In addition, numerous vertical and angle cores have been collected to examine

changes to the iron surface over time and to evaluate the formation of secondary precipitates which may

affect wall performance over time. Coring was done vertically (perpendicular to ground surface) and on an

angle (30E). The former method provided continuous vertical iron cores, while the latter provided a

transverse core through the wall with the aquifer-iron interfaces intact (front and back of the wall). These

cores continue to be under study. Inorganic carbon contents increase dramatically at the up-gradient aquifer

sediment-iron interface and decrease within the wall moving down gradient, reaching background levels

within 10 cm down gradient from the down gradient iron-aquifer sediment interface. Total inorganic carbon

content has increased over time within the wall. 

Results of geochemical sampling on site indicate that iron corrosion is proceeding within the wall. There

are significant reductions in Eh (to >-400 mv), increase in pH (to >10), absence of DO, and decrease in

alkalinity. Down gradient of the wall (1.5 m), pH returns to near neutral and Eh is quite variable with depth.

Over time there have been indications that a redox front is slowly migrating down gradient within the first

few meters of the wall. Water levels indicate little difference (<0.1 m) between wells completed and

screened at similar depths up gradient and down gradient of the wall, indicating that the wall continues to

effectively function as a “permeable” reactive barrier.

Sampling results for chromate indicate that all chromate was removed from the groundwater within the first

15 cm of the wall as expected. No chromate is detected down gradient of the wall either in the multilayer

samplers or in the 5-cm compliance wells located immediately behind the wall.

The vast majority of the multi-layer sampling ports show reduction of the chlorinated compound

concentrations to less than regulatory target levels. Only one port (ML25-1) continues to show levels above

target concentrations. This is the deepest port in the middle of the wall where the organic solvent compound

concentrations are highest.
Treatment Walls NATO/CCMS Pilot Study Phase III

5

Other Inorganic Contaminants

Other inorganic contaminants currently under study include nitrate and arsenic. These inorganic

contaminants are transformed via oxidation-reduction reactions to less mobile and/or less toxic forms and

immobilized as insoluble precipitates or through essentially irreversible adsorption reactions. Zero-valent

iron can effectively transform nitrate to reduced forms as confirmed with mass balance determinations. The

use of different organic materials also further reduces the nitrate and experiments are in progress to account

for all nitrogen forms produced. Experiments have also explored the use of iron-organic material mixtures.

Experiments with arsenic are currently underway and we are exploring conversion to reduced and oxidized

forms for immobilization in a reactive barrier. 

Disclaimer

Although the research described in this article has been funded wholly or in part by the United States

Environmental Protection Agency, it has not been subjected to the Agency’s peer and administrative review

and therefore may not necessarily reflect the views of the Agency and no official endorsement may be

inferred.

Discussion

Zero-valent iron walls can affect the pH of chlorinated solvent plumes. The pH of a plume can increase to

10 to 11 after passing through a 100% iron wall. This change is always observed due to the nature of the

corrosion reaction that takes place and occurs regardless of whether the plume contains chlorinated solvents

or chromium. However, the iron can be mixed with pyrite or native aquifer sediments (e.g., 50% iron and

50% sediment) to buffer the pH in the walls to less than 7.5. The buffered walls exhibit an environment

favorable for microbial activity, so there is a lot of sulfide is produced. Additionally, hydrogen is produced

during the corrosion reaction at the iron surface. The production of hydrogen is significant—up to 1,000 nM

of dissolved hydrogen has been measured at the Elizabeth City site.

Puls’s data documented a buildup of chromium (2% by weight after one year) in the PRB installed at

Elizabeth City, and Puls expects it to increase in the future. This expectation is based on two-year column

tests in which over 3,000 pore volumes were pumped through under accelerated conditions. Although a

buildup occurred that reduced the permeability of the column, it did not impact the reduction of chromate

in the groundwater. The pressure buildup caused by the precipitate occurred after pumping approximately

500 pore volumes. It is difficult to extrapolate these laboratory results to field-scale tests, and Puls said that

they will continue to monitor for long-term reactivity in the wall.

The residence time for pore water in the fastest part of the barrier is two days. The hydraulic conductivity

of the aquifer varies over its 24-foot thickness, and the highest hydraulic conductivity occurs within a 2-

meter interval. This interval contains the highest concentrations of Cr , but most of the TCE is found in 6+

deeper zones. The reduction of chromium causes no inhibitory or competitive effect on the degradation of

TCE, because the chromium and TCE are present in different intervals. While the hydraulic conductivity

of the aquifer ranges from 2-5 inches/day, the hydraulic conductivity of the wall is 10-12 inches/day. 
Treatment Walls NATO/CCMS Pilot Study Phase III

 WCI Umwelttechnik GmbH, Sophie Charlotten - Str.33, 14059 Berlin, Germany, tel: 49/30-3260-9481, fax: 49/30- 1

321-9472

6

Technical Construction of Treatment Walls

Permeable Treatment Walls—Design, Construction, and Cost

Eberhard Beitinger1

Abstract

The Design of Permeable Reactive Barriers (PRB) is discussed with reference to preventing effects such as

inflow of fine soil particles, precipitation of carbonates, iron and manganese, and loss of effectiveness by

uncontrolled growth of biomass.

Several construction methods are presented and evaluated in terms of design objectives, installation

methods, implementation, effectiveness and cost.

1. In Situ Groundwater Remediation with Permeable Reactive Barriers

According to EPA, “...a permeable reactive barrier (PRB) is a passive in situ treatment zone of reactive

material that degrades or immobilizes contaminants as groundwater flows through it. Natural gradients

transport contaminants through strategically placed treatment media. The media degrade, sorb, precipitate,

or remove dissolved organics, metals, radionuclides, and other pollutants. These barriers may contain

reactants for degrading volatile organics, chelators for immobilizing metals, nutrients and oxygen for

microorganisms to enhance bioremediation, or other agents.” [1]

The following report is focused on construction methods that prevent negative effects that may reduce the

effectiveness of the treatment media, negatively change the hydraulic conditions, or reduce the long-term

PRB performance. This includes the application of knowledge and experience from groundwater well design

and the evaluation of innovative technologies to remove and/or reactivate the treatment media in situ.

Permeable Reactive Barriers have commonly been, or are designed to be, installed through excavation and

replacement. This method is limited to approximately 8 m in depth, and the costs becomes prohibitive at

greater depths. Alternative installation methods for greater depths include slurry wall construction, high

pressure jetting, deep soil mixing, and hydro fracturing in the United States [2]. In Germany large diameter

vertical borings, slurry methods and a modified deep wall construction are under evaluation or investigation,

but have not yet been tested in the field.

2. Design Objectives for Permeable Treatment Walls

The hydraulic behaviors of the two major permeable treatment wall design types—funnel and gate systems

or continuous reactive walls—are based on the hydraulic permeability of the whole construction system,

including permeability of filter layers, screens and the treatment media itself. The system permeability of

the wall construction should be at least twice the permeability of the aquifer. It might be better to choose

a factor of 10 times higher permeability for the wall system because of to all the limiting parameters that will

reduce permeability with time. The major limiting factors are expected to be as follows:

 • inflow and settling of fine-grained soil particles, which will block the pore spaces and reduce the

permeability;
Treatment Walls NATO/CCMS Pilot Study Phase III

7

 • precipitation of carbonates such as calcium or magnesium carbonates, iron oxides/hydroxides, and

ferrous carbonate or other metal precipitates in the filter layer or treatment media;

 • uncontrolled growth of microorganism, such as bio-clogging; and

 • other, mainly long-term and unknown effects, which may reduce the permeability.

The following design objectives are focused mainly on the use of activated carbon as the treatment media.

But most of these design objectives may also be useful if zero-valent iron (ZVI) or other media are chosen

as reactive filling materials.

Major design objectives are to:

 • Prevent the inflow of fine-grained soil particles by installing a filter layer between the adjacent soil and

the treatment media according to the well-known filter criteria;

 • Prevent changes to the physical and chemical properties of the groundwater such as temperature,

pressure, pH, oxygen content (redox potential), nutrients, or be aware that these changes will result in

precipitation or microorganism activity that might be of advantage in specific cases;

 • Design a wall-system that allows the removal and replacement of the treatment media after a period of

several years or decades;

 • Design a pipe-system in the wall to allow the injection of water or air for flushing to eliminate

precipitates or sludges or to remix the media by turbulence; and

 • Design openings for inspection, removal or sampling of the treatment media.

The above list of design objectives may not be applicable in every case. Based on the specific site

conditions, the nature of contamination and the remediation goals, additional design objectives have to be

considered.

3. Design Criteria

The design of Permeable Reactive Barriers needs to carefully address the following design criteria. In

general, investigations to obtain design data should be planned with reference to the specific design

parameters for a long-term performance of the underground treatment wall.

Design criteria are:

 • Geology

 • Hydrology

 • Depth to groundwater

 • Thickness of the aquifer

 • Groundwater flow direction and gradient

 • Permeability

 • Grain-size distribution

 • Pore volume

 • Groundwater chemistry, such as pH-value, redox-potential, electrical conductivity, oxygen content,

temperature, hardness, iron and manganese concentration, sulfates, and nutrients

 • Biological activity (microorganisms)

 • Pollutants (concentration, spread, transfer conditions)

 • Site conditions, topography, and access

 • Sensitivity of aquifer regarding potential users

 • Damage to existing buildings regarding settling problems

 • Available time for remediation

 • Effectiveness of treatment
Treatment Walls NATO/CCMS Pilot Study Phase III

8

Figure 1. Peat-filled trench [1]

Figure 2. Solid free trench [2]

 • Costs

 • Monitoring.

4. Construction Methods

During the last five years, a variety of construction methods have been developed in Canada, the United

States and in Europe. Some of these construction methods have been implemented or tested in field scale.

Others are proposed for implementation but have not yet been adapted.

4.1 Peat-Filled Trench

The University of South Carolina [1] obtained a patent October 15, 1991, for a method for in situ removal

of hydrocarbon contaminants from groundwater using a permeable barrier comprising a peat material or,

an immobilized nutrient layer and peat material layer in series.

As shown in Figure 1, it is assumed that an open trench would be excavated and filled with the reactive

media. The construction method is simple and inexpensive.

The specific costs for open trenches are mainly

the costs for excavation and disposal, including

costs for lowering groundwater and, if necessary, stabilizing the open ditch.

The actual costs, without stabilization, may

range between US$10-100/m2. Costs for stabilization will vary within US$50-100/m2.

4.2 Solid-Free Trench

In February 1992, a German patent [2] was granted to

Dr. Haldenwang and Dr. Eichhorn for a solid-free

trench construction in which the walls of an open

trench are permeable and stabilized and the

groundwater flows through it (Figure 2). Additionally,

piping for aeration or circulation of water within the

open trench was proposed including measures for

capping and emission control. 

For field testing, a 6 m deep, 1.4 m wide and approximately 10 m long trench was established on a refinery

site and tested for several years.

No solid media is used as reactive media but

according to its function as an underground permeable

barrier this technology is mentioned as an alternative

within the group of permeable reactive barriers. To

stabilise the wall, permeable sheet piles have been

proposed. The specific costs are estimated to range

between US$200-400/m2 not including any preparation or other aeration/de-aeration installations.
Treatment Walls NATO/CCMS Pilot Study Phase III

9

Figure 3. Open trench construction of Permeable

Reactive Barriers

Figure 4. Permeable Treatment Wall Construction

Figure 5. Cross-Section—Treatment Wall

4.3 Open Trench Construction

In several cases, open trenches filled with reactive

media have been planned or executed. Vertical sheet

piling was used to stabilise the open excavation

ditch and pulled after filling of reactive media was

completed. The maximum width of the reactive wall

is dependent on the distance that the sheet pile walls

are driven into the ground (Figure 3).

One of the major disadvantages is that there is no

filtration layer between the adjacent soils and the

reactive media to help avoid inflow of fine-grained soil particles. Also recovering the filling material

(reactive media) is not possible without destroying the structure itself. The specific construction costs will

vary (without reactive media) in the range of US$100-200/m2.

4.4 Treatment Wall Construction by WCIUmwelttechnik GmbH

WCI Umwelttechnik GmbH has developed a special

method of wall construction that allows the

adsorbing or other reactive media to be recovered

without the need to demolish and rebuild the wall

structure [3]. The patented system includes filtration

layers to prevent inflow of fine soil particles and

measures to avoid precipitation by oxidation of iron

and manganese. As shown in Figure 4, the main

design components are:

 • a filter layer, consisting of gravel or sand pack,

located between the trenched aquifer and the

interior wall elements;

 • interior wall elements, made of specially

designed brick elements or pre-cast concrete

shells to define an interior filling space for the

treatment media, including openings for the

groundwater through-flow and horizontal

elements to stabilise the earth pressure;

 • a clay seal to prevent inflow of stormwater and

contact with atmospheric oxygen;

 • a cover plate as an opening device for the

recovery and replacement of the treatment media

—additionally these covers can be made

waterproof and airtight.

Some typical dimensions are shown for the cross

section of the wall construction in Figure 5.
Treatment Walls NATO/CCMS Pilot Study Phase III

10

Figure 6. Construction Phase 1 Figure 7. Construction Phase 2

Depending on the usually low groundwater flow velocity, the retention time of groundwater in a 0.30 m

thick treatment media layer is between several hours and one or two days.

In order to dimension the retention time and media thickness, new approaches must be developed based on

test results in the field. In particular for the use of activated carbon, the retention time, surface capacity and

the height or width (thickness) of media layers in the permeable treatment walls will differ widely from

those dimensioning criteria used in conventional aboveground activated carbon filters (Table 1).

Table 1: Dimensioning Criteria

Dimensioning Criteria PRB-adsorbing Conventional

Activated Carbon Filter

Retention time 5 - 30 hrs 1 - 2 hrs

Surface load capacity 0.02 - 0.06 m/h 10 - 15 m/h

Height/width filter layer 0.3 - 0.5 m 2 m

Installation methods

The installation methods are described in detail in Figures 6 to 11.

In phase 1 (Figure 6) sheet pile walls will be driven into the earth. The distance between the piles will be

1 m or more, according to the chosen system width. The sheet piles will be installed 2 m into the bottom

layer of the aquifer.

In phase 2 (Figure 7) excavation will commence within the two sheet pile walls. An open ditch will be used

to lower the groundwater table below the excavation level.

After completion of the excavation (Figure 8), a concrete base layer will be cast in situ and dewatering will

be continued. The horizontal earth pressure will be supported with beams.

Now the installation of the forms and placing the filter gravel can begin on the concrete layer (Figure 9).

As the construction grows, groundwater level can be allowed to rise again.
Treatment Walls NATO/CCMS Pilot Study Phase III

11

Figure 8. Construction Phase 3

Figure 9. Construction Phase 4

Figure 10. Construction Phase 6 Figure 11. Construction Phase 6

As the wall construction continues (Fig. 10), sheet

piles can be raised accordingly and the treatment

media will be filled into the interior space of the

forms.

Finally, the wall cover, the clay seal and some other

measures can be installed to complete the

construction (Figure 11). In addition, bank

protection (if located near a shore as in the given

example), monitoring wells and a roadway for the

tank trucks (to remove and replace the treatment

media) may be installed.
Treatment Walls NATO/CCMS Pilot Study Phase III

12

Figure 12. Construction Progress in Longitudinal Direction

Figure 13. Dr.-Ing. Steffen—Barrier Construction Phases

Figure 12 shows an approximately 30-m long constructed longitudinal section. All the necessary installation

phases from excavation to covering the completed wall are illustrated. For longer walls, several construction

sections may be useful.

The construction costs for permeable treatment walls with a depth of 8 m are estimated to range from DM

850 to 1400 per square meter (US$500-900/m2) without the treatment media (an additional DM 300/m2

[US$180/m2] for an activated carbon layer of 0.3 m thickness).

4.5 Construction Patented by Dr. Steffen Ingenieurgesellschaft

Another construction method based on a slurry

wall technology was patented (June 27, 1992)

by a German consultant, Dr.-Ing. Steffen

Ingenieurgesellschaft mbH, Essen [4]. The

patented idea is to press a sheet pile or a

double-T-formed steel beam down into the

ground and to inject activated carbon by

pulling the beam out again (Figure 13). The

overlapping process of pressing and injection

will form a continuous wall of reactive media

with a width of approximately 0.10-0.15 m.

Depending on geology, depths to 20 m or more

can be reached. When the adsorbent capacity

has been exhausted, a second wall is planned

upstream.

There is no filter layer and the recovery of the media is not anticipated. Such a wall construction may be

blocked easily by inflow of fine soil particles or by precipitation effects. The cost range is estimated to be

US$200/m2 without the media itself.
Treatment Walls NATO/CCMS Pilot Study Phase III

13

Figure 14. Large Diameter Borehole

Construction Method

Figure 15. Sax+Klee GmbH, Germany

Figure 16. Funnel and Gate System,

University of Waterloo

4.6 Large Diameter Borehole Construction Method

This construction method is used in the U.S. as well as in Germany. Mull und Partner, Hannover, Germany,

has proposed a continuous overlapping iron-filled wall in Nordrhein-Westfalen to prevent downstream

contamination of chlorinated hydrocarbons emitted by a laundry.

Usually, borehole diameters from 0.8-1.6 m are

chosen. Individually located bore holes, established

in several lines are known as an “iron-fence.” The

method is applicable for continuous walls, if the

bore holes overlap, or for gate constructions. Depths

to 20-25 m may be reached with no problems. By

overlapping the bore holes, a loss of approximately

15-20 % of the filling material has to be allowed.

Depending on the geology, in most cases steel pipes

have to be used for stability of the borehole during

excavation. This construction method is shown in

Figure 14.

The costs for this method are US$300-500/m2,

excluding filling materials.

4.7 Method developed by Sax+Klee, Germany

The German construction company, Sax+Klee

GmbH, Mannheim, has developed a modified large

diameter borehole construction method for a funneland-gate PRB system. The innovative approach

includes using bore holes filled with impermeable

clay to connect the funnel sheet pile walls with the

gate structure. Also new is the proposed installation

of an interior well screen with four wings to create

a filler zone between the surrounding soils and the

reactive media (Figure 15).

The costs for a 40 m wide and 19 m deep gate

construction are estimated to be approximately

US$1,400/m2.

A filling material of petroleum activated carbon is

proposed for a site with groundwater contaminated

with petroleum hydrocarbons and aromatics

(BTEX). The site is a former U.S. Army tank farm.

4.8 “Funnel and Gate” Patented by University of

Waterloo

The funnel and gate method (Figure 16) has been

patented worldwide since April 23, 1992, for the

University of Waterloo, Canada [5].
Treatment Walls NATO/CCMS Pilot Study Phase III

14

“Contaminated groundwater is treated in situ, by funnelling the water through a gate or gates in a watertight

in-ground wall. Treatment material in the gate breaks down the contaminant, or otherwise removes the

contaminants from the flowing water. A removable caisson is first driven into the ground, excavated, and

then a receptacle, for the treatment material, is lowered into the hollow interior.”

The patent itself shows a wide variety of construction elements, such as caissons made of steel (receptacles),

slurry and sheet pile waterproof walls, removable baskets in the receptacle, removable caissons, etc.

Costs for the construction methods proposed by the University of Waterloo may vary in the range of

US$200-1,000/m2.

4.9 Bio-Polymer Slurry Trenching Technique

Geo-Con, a subsidiary of Woodward-Clyde, reported (December 23, 1997) the performance of a funnel-andgate project for the U.S. Department of Energy, Oak Ridge, as follows:

Geo-Con recently completed what we believe to be the first permeable reactive wall installed using

the bio-polymer slurry trenching technique. The reactive gate consists of a 30-foot (900 cm) iron

filing “gate,” with two 100-foot (30 m) wings composed of pea gravel. The system is 2 feet (60 cm)

wide and extends down to bedrock, which is approximately 26 feet (780 cm) below ground surface.

The purpose of the permeable reactive wall is to collect a plume of groundwater contaminated with volatile

organic compounds (VOCs). The trench intercepts the groundwater, funneling it through the pea gravel to

the iron filing gate. As the groundwater passes through the filing gate, the VOCs are removed.

This project was originally contracted to another company, to be built by excavating dry and backfilling with

the filings and pea gravel. Upon nearing completion of the first cut-approximately 20 feet (600 cm) deep,

the trench collapsed. Geo-Con proposed to complete the project using a bio-polymer slurry to maintain

trench stability during gravel and iron filling installation. Geo-Con was able to complete the project in three

days, expediting the project schedule and providing a considerable cost saving over other installation

methods. [6]

The author estimates costs for this technique to be in the range of US$300-500/m2, not including the filling

materials.

4.10 Other Construction Methods

The following listing of other construction methods under discussion or developed for the installation of

Permeable Reactive Barriers is not complete; it reflects the knowledge of the author about actual

information.

 • Enviro Wall is a trademark of Barrier Member Containment Co. Ltd. and Argonne National Laboratory.

By help of a guide box installation, an HDPE geomembrane will be installed, and a groundwater

collection and redistribution system will guide groundwater to a pass through module. The open trench

will be refilled with site soil and capped with bentonite after installation of membranes and horizontal

collection systems.

 • The PRB Action Team reports studies using high pressure jet grouting systems to create either

permeable grout diaphragms or columns [7]. Overlapping these structures creates longer, continuous

barriers. In a field study, two grout formulas have been tried: a resin/granular cast iron slurry and a
Treatment Walls NATO/CCMS Pilot Study Phase III

15

kaolinite clay/iron slurry. Also, deep soil mixing and hydrofracturing (horizontal and vertical) are under

discussion in the U.S. as alternative installation methods.

 • Envirotreat Ltd., UK, is currently working on the development of active containment systems using

special clay materials. Addition of reactive materials to clay barriers is also under investigation.

Overlapping large diameter bore holes are one of the construction methods is considered to be

applicable.

 • Several German engineering companies have proposed the installation of gates in the form of concrete

structures for horizontal or vertical through-flow (Peschla und Rochmes, Edenkoben, and WCI

Umwelttechnik, Essen). These are conventional underground concrete structures, caissons, or chambers

that contain removable reactive media. These buildings are expensive and limited to use as single gate

structures.

 • WCI-Umwelttechnik GmbH is working on a deep wall construction method using modified slurry wall

techniques with additional caissons as containment for the reactive media.

Finally, the injection of reactive material into the unsaturated or saturated zone, not forming a defined

structure, is a potential method, but not discussed in this report.

5. Costs

Table 2 presents a rough overview to costs for the several construction techniques mentioned. They do not

include the filling material. They also depend on national markets and on actual/future experience and

developments. Some of the actually methods may disappear in the near future, and others may show

economical advantages. Generally, we need more field testing and full-scale experience than we now have.

Table 2: Cost Overview

Construction method Max Cost Range

Depth [m] [US$/m2]

4.1 Peat-filled trench 5 10 -100

4.2 Solid-free trench 10 200 - 400

4.3 Open trench construction 10 100 - 200

4.4 WCI-treatment wall 12 500 - 900

4.5 Dr. Steffen’s patent 20 200

4.6 Large diameter borehole 20-25 200 - 500

4.7 Sax+Klee system 20 1000 - 1500

4.8 Funnel and Gate 20 200 - 1000

4.9 Bio-polymer slurry 30 300 - 500

4.10 Other construction methods 10 100 - 300

20 200 - 1000

30 500 - 1500

6. Conclusions

In our estimation, it is too early to evaluate construction methods in general. They strongly depend on the

required depth, hydrogeology, contaminants, remediation targets, etc. Most of the construction methods

discussed are unproven in relation to PRBs, or are under testing.
Treatment Walls NATO/CCMS Pilot Study Phase III

16

A wide variety of construction elements are available. The future will show which systems are most

applicable to shallow and deep aquifers and that are economically sound.

Full-scale applications and field tests will be helpful in establishing the innovative PRB technology as a

proven and available remediation alternative. Besides the research and development activities regarding

effective reactive media, we have to deliver inexpensive, durable construction techniques to install

(recoverable) reactive media in the contaminated groundwater plume.

7. References

[1] United States Patent Number 5,057,0227, Cohen, University of South Carolina, Columbia, S.C.,

October 15, 1991.

[2] Patentschrift DE 41 05 987 C2, Haldenwang, Eichhorn, February 26, 1991 “Vorrichtung und Verfahren

zur in situ-Behandlung verunreinigter Grundwässer”.

[3] Patentschrift DE 44 25 061 C1, WCI-Umwelttechnik GmbH, July 15, 1994 “Permeables

Behandlungsbett zur Reinigung kontaminierter Grundwasserströme in situ”.

[4] Patentschrift DE 4221 198 C2, Dr.-Ing. Steffen Ingenieurgesellschaft mbH, June 27, 1992 “Verfahren

zum Entfernen von wasserlöslichen sorbierbaren Schadstoffen aus einem abströmenden Grundwasser

in der Umgebung eines Kontaminationsherdes”.

[5] International Patent W093/22241, Cherry, Vales, Gillham, University of Waterloo, Canada, April 23,

1992 “System for Treating Polluted Groundwater”.

[6] Geo-Con, internal company information, not published, December 1997.

[7] Permeable Reactive Barrier Action Team, Summary of the Remediation Technologies Development

Forum, Virginia Beach, Virginia September 18-19, 1997.

Discussion

It was suggested that treatment walls may not need to intercept all of the contaminated groundwater, due

to the expense of constructing the walls, and could be designed to treat (for example) only 90% of the

groundwater flow. This would make the technology cheaper and more appealing. Beitinger agreed that such

walls can be constructed (e.g., an iron fence or non-overlapping borehole reactors), but the decision is

usually based on whether the regulating authority would permit continued flow of untreated water.
Treatment Walls NATO/CCMS Pilot Study Phase III

 Submitted for publication in: Proceedings, American Society of Civil Engineers, 1998. 1

 University of Waterloo, Department of Earth Sciences, 200 University Avenue W., Waterloo, ON N2L 3G1, Canada, 2

tel: +519-888-4658, fax: +519-746-7484, e-mail: rwgillha@sciborg.uwaterloo.ca

17

Development of Iron-Based Reactive Barrier

Technologies for Remediation of Chlorinated Organic

Contaminants in Groundwater1

Robert W. Gillham2

Abstract

Granular iron was not recognized as an effective reductant for promoting the dechlorination of halocarbons

in aqueous solution until the late 1980s. Furthermore, the suggestion that granular iron be used, for in situ

remediation of groundwater containing halocarbons was initially met with a high degree of skepticism. In

the intervening years, the use of granular iron for groundwater remediation has emerged as a new and

significant environmental technology. This paper outlines some of the major contributions that have led to

the growing scientific and market acceptance of the technology.

Introduction

Recognizing the limitations of pump-and-treat methods for groundwater remediation (NAS, 1994),

considerable attention has now turned to the use of in situ permeable reaction barriers (PRBs). An early

review of this topic (prepared in 1992) is given in Gillham and Burris (1997). In this concept, a permeable

“wall” containing the appropriate reactants is constructed across the path of a contaminant plume. As the

contaminants pass through the reactive material under passive groundwater flow, they are removed by

chemical and/or physical processes. Proponents of PRBs cite reduced capital cost, low operating and

maintenance costs, and conservation of water and energy among the potential advantages over pump-andtreat systems.

The use of granular iron for removal of chlorinated organic contaminants has led the interest in PRBs. With

initial recognition in 1989, this concept is now emerging as a significant technology for groundwater

remediation. This paper summarizes the major steps in both the scientific understanding and acceptance,

and in the commercial development of the technology.

Scientific Development

In studies of the effect of sampling-well materials on sample integrity, (Reynolds et al., 1990) it was

observed that the concentration in aqueous solution of several chlorinated organic contaminants declined

when in contact with certain metals. Though this observation was made in 1984, it was not until 1989 that

the potential significance with respect to groundwater remediation was recognized. Further testing

confirmed the earlier results and revealed various aspects of the process. In particular, based on our early

understanding of the reactions, the passive in situ use of granular iron was proposed as an alternative for

remediation of groundwater containing halocarbons. Response to publication of the early experimental

results and concepts (Gillham and O’Hannesin, 1992) generally varied from skepticism to disbelief.

A degree of credibility and the attention of the scientific community was gained through two papers

published in 1994. Gillham and O’Hannesin (1994) showed a wide range of contaminants to degrade, in

the presence of granular iron, at rates that are several orders of magnitude greater than natural abiotic

degradation rates reported in the literature. Evidence was presented to indicate that the reaction was abiotic
Treatment Walls NATO/CCMS Pilot Study Phase III

18

reductive dechlorination and followed pseudo-first order kinetics. Matheson and Tratnyek (1994), from

detailed studies of the degradation of carbon tetrachloride, showed results consistent with those of Gillham

and O’Hannesin (1994), and further proposed three possible reaction mechanisms. All involved sequential

dechlorination through one-electron transfer. Both Matheson and Tratnyek (1994) and Gillham and

O’Hannesin (1994) proposed two simultaneous redox reactions 1) oxidation of the iron coupled with

reduction of water and 2) oxidation of iron coupled with reduction of the chlorinated organic compound.

Reaction (1) results in the production of H2 and Fe and an increase in pH as a consequence of the 2+

production of OH . Reaction (2) causes a further release of Fe , the release of Cl and production of a less - 2+ -

chlorinated organic compound. Based on generally low concentrations of dissolved iron observed in batch

and column tests, it was proposed that iron precipitates as Fe(OH) 2

, limiting the pH increase to values in the

range of about 9 to 10.

The scientific aspects of the technology “came of age” through an American Chemical Society symposium

“Contaminant Remediation with Zero-Valent Metals” organized by Drs. Paul Tratnyek and Martine

Reinhard (209th ACS National Meeting, Anaheim, California, April 2-7, 1995). As an indication of the

increasing interest, over 40 abstracts were submitted for presentation. Information was presented indicating

that the reactions occur on the solid surfaces rather than in free solution (Weber, 1995 and Totten and

Roberts, 1995); based on relatively low production of intermediate degradation products, the one-electron

sequential degradation pathway was brought into question (Sivavec and Homey, 1995 and Orth and

Gillham, 1995). Totten and Roberts (1995) showed that two-electron transfers can occur, and through this

and subsequent work by Roberts et al. (1996) and others, it is now widely accepted that at least for the

chlorinated ethenes, trichloroethene and tetrachlorothene, there are two competing pathways, single- electron

transfer with DCE isomers and vinyl chloride as intermediate products and two-electron transfer where

chloroacetylene is the intermediate. Though both pathways are followed simultaneously, in most situations

it appears that the latter is dominant. This is favorable of course since chloroacetylene is highly unstable,

while the DCE isomers and vinyl chloride generally degrade more slowly than their parent compounds (TCE

and PCE).

Other important contributions concern the affects of inorganic constituents (other than FeE). Sivavec and

Horney (1995) noted that the iron surfaces are generally covered with an iron oxyhydroxide coating,

suggesting a two-step degradation process, sorption onto the surface followed by electron transfer. Noting

that electron transfer must occur through the surface coatings, the nature of the coatings and their ability to

conduct electrons becomes an important issue. Precipitation of Fe(OH) , a consequence of both reactions 2

(1) and (2) as presented previously, could form a protective coating thus resulting in a loss of activity over

time. In subsequent studies, however, (Odziemkowski et al., 1998) it has been shown that Fe(OH) is 2 

converted to magnetite, which is conducting to electrons, thus maintaining the activity of the iron surfaces.

It is also recognized that the increase in pH caused by the reduction of water can have a profound influence

on the inorganic chemistry of natural waters. In particular, in response to increasing pH, bicarbonate in

solution is converted to carbonate, which can result in the precipitation of carbonate minerals such as siderite

and calcium carbonate. These have the potential to form surface coatings that could reduce activity and

ultimately clog pores, resulting in loss of permeability. In pilot-scale laboratory tests reported by Mackenzie

et al. (1995), there was no apparent loss of reactivity of the iron, though pore blockage was identified as an

issue requiring further investigation. A further contribution of the 1995 ACS symposium was the

identification of materials with much higher reaction rates than iron. For example, Korte et al. (1995), and

later Liang et al. (1997), showed that a bimetallic consisting of palladium plated onto the iron surface could

increase reaction rates, by up to two orders of magnitude.

Current areas of active investigation include fundamental studies of reaction mechanisms and pathways, and

degradation rates and pathways for a wider range of organic contaminants. Work continues to identity

effective bimetal couples and concerns remain regarding the long-term integrity of the catalytic
Treatment Walls NATO/CCMS Pilot Study Phase III

19

enhancement. Certainly the most pressing practical issue is the long-term performance of zero-valent iron

in natural subsurface environments. Various research groups are continuing to investigate the changes over

time in the surface characteristics of iron in contact with various types of groundwater, and the nature and

consequences of precipitates that may form.

Commercial Development

As discussed in NAS (1997) and summarized in Macdonald (1997), many companies developed for the

purpose of marketing new environmental technologies have had a remarkably poor record of success. In

particular, the requirement for regulatory acceptance of new technologies and the fact that the incentives

for site owners generally encourage the delay of remedial action are significant contributors to a sluggish

environmental market. The in situ iron technology is marketed by EnviroMetal Technologies Inc. (ETI)

under a licensing agreement with the University of Waterloo. ETI was incorporated in 1992, at a time when

there was a high degree of skepticism. In addition, because capital costs of installing an in situ iron treatment

system are not particularly advantageous, the main financial incentive is in greatly reduced long-term

operating and maintenance costs. Clearly, in 1992, it was not possible to demonstrate these savings.

The first important contribution to market access was the development, in 1992, of an in situ demonstration

at Canadian Forces Base Borden. The early results of that test showed that the dechlorination reaction would

indeed proceed in situ, and each passing year of consistent performance, raised the level of confidence in

the efficacy of the technology. As reported in O’Hannesin and Gillham (1998) this facility performed

consistently and effectively over the five- year duration of the test. The second major contribution to market

development was the first in situ treatment system at a commercial site Yamane et al., 1995). This occurred

in late 1994 and early 1995, at a time when general knowledge of the technology was at a low level and

long-term in situ performance remained as a significant unresolved question. The interest of the client in

new technologies, and willingness to support the testing required to develop the design was critical to this

initial application, and a significant contribution to market development.

Another important development was the identification of large quantities of granular iron available at

reasonable cost. Currently there are three major suppliers, all of whom recover cuttings from metal

machining and fabrication operations, treat and grade the materials and resell the product for a variety of

commercial uses. These materials have proven to be highly effective as the reductant in the dechlorination

process.

While regulatory concerns can slow technology implementation, regulators, particularly the EPA, have

contributed in a major way to the growing acceptance of the iron technology. In addition to favorable results

reported by EPA researchers, an above-ground demonstration installed in New Jersey in 1995 was

monitored under the EPA SITE program, resulting in a favorable report (US EPA, 1997). Similarly, an in

situ demonstration in New York State promises to result in a favorable report under the SITE program

(anticipated in 1998). Inclusion of PRBs, as a Working Group under the Remediation Technology

Development Forum has also played an important role in disseminating information concerning PRBs

generally and the use of granular iron in particular.

ETI currently reports the installation of ten demonstration facilities and twelve full-scale treatment facilities.

While the numbers remain small, the trends suggest a growing awareness and acceptance of the technology.

In particular, all recent installations have been full-scale, the time between initial contact and

implementation is decreasing and the number of inquiries continues to increase.
Treatment Walls NATO/CCMS Pilot Study Phase III

20

References

Gillham, R.W. and O’Hannesin, S.F., 1992. Metal-catalyzed abiotic degradation of halogenated organic

compounds. IAH Conference: In Modern trends in hydrogeology, Hamilton, Ontario, May 10-13, pp.

94-103.

Gillham, R.W. and O’Hannesin, S.F., 1994. Enhanced degradation of halogenated aliphatics by zero-valent

iron. Ground Water, Vol.32, No.6, pp. 958-967.

Gillham, R.W. and Burris, D.R., 1997. Recent developments in permeable in situ treatment walls for

remediation of contaminated groundwater. Chapter 21: Subsurface Restoration, Eds. Ward, C.H.,

Cherry, J.A. and Scalf, M.R., Ann Arbor Press, Inc., Chelsea, Michigan, pp 343-356.

Liang, L., Korte, N., Goodlaxson, J.D., Clausen, J., Fernando, Q. and Muftikian, R., 1997. Byproduct

formation during the reduction of TCE by zero-valent iron and palladized iron. Ground Water

Monitoring and Remed., Winter, pp.122-127.

Macdonald, J.A., 1997. Hard times for innovative cleanup technology. Environmental Science and

Technology, Vol.31, No.12. pp.560-563.

Mackenzie, P.D., Baghel, S.S., Eyholt, G.R., Horney, D.P., Salvo, J.J. and Sivavec, T.M., 1995. Pilot-scale

demonstration of reductive dechlorination of chlorinated ethenes by iron metal. In proceedings of 209th

American Chemical Society National Meeting, Anaheim, CA, Vol.35, No.1, pp.796-799.

Matheson, L.J. and P.G. Tratnyek, 1994. Reductive dehalogenation of chlorinated methanes by iron metal.

Environ. Sci. Technol., Vol.28, No.12, pp. 2045-2053.

National Academy of Sciences, 1994. Alternatives for ground water cleanup. Report of the National

Academy of Science Committee on Ground Water Cleanup Alternative. Washington, D.C., National

Academy Press.

National Academy of Sciences, 1997. Innovations in Groundwater and soil cleanup: From concept to

commercialization. National Research Council,

Committee on Innovative Remediation Technology, National Academy Press, Washington, D.C.

Odziemikowski, M.S., T.T. Schuhmacher, R.W. Gillham, and E.J. Reardon, 1998. Oxide Film Formation

on Iron in Simulating Groundwater Solutions: Raman Spectral and Electrochemical Studies. Corrosion

Studies (in press).

O’Hannesin, S.F. and Gillham, R.W., 1998. Long-term performance on an m situ “iron wall” for

remediation of VOCs. Ground Water, Vol.36, No.1, pp. 164-170.

Orth, S. and Gillham, R.W., 1996. Dechlorination of trichloroethene in aqueous solution using Fe .

0

Environmental Sci. and Technol., Vol.30, No.1, pp.66-71.

Reynolds, G.W., Hoff, J.T. and Gillham, R.W., 1990. Sampling bias caused by materials used to monitor

halocarbons in groundwater. Environmental Science and Technology, Vol.24, No.1, pp.135-142.
Treatment Walls NATO/CCMS Pilot Study Phase III

21

Roberts, A.L, Totten, L.A., Arnold, W.A., Burris, D.R., and Campbell, T.i, 1996. Reductive elimination of

chlorinated ethylenes by zero-valent metals. Environmental Science and Technology, Vol. 30, No.8,

pp.2654-2659.

Sivavec, T.M. and Horney, D.P., 1995. Reductive dechlorination of chlorinated ethenes by iron. In

proceedings of 209th American Chemical Society National Meeting, Anaheim, CA, Vol.35, No.1,

pp.695-698.

Totten, L.A. and Roberts, A. L., 1995. Investigating electron transfer pathways during reductive

dehalogenation reactions promoted by zero-valent metals. In proceedings of 209 American Chemical th

Society National Meeting, Anaheim, CA, Vol.35, No.1, pp.706-708.

Weber, E.J., 1995. Iron-media reductive transformations: Investigation of reaction mechanism. In

proceedings of 209 American Chemical Society National Meeting, Anaheim, CA, Vol. 35, No.1, th

pp.702-705.

Yamane, C.L, Warner, S.D., Gallinatti, J.D., Szerdy, F.S., Delfino, T.A., Hankins, D.A., Vogan, J.L., 1995.

Installation of a subsurface groundwater treatment wall composed of granular zero-valent iron. In

proceedings of 209 American Chemical Society National Meeting, Anaheim, CA, Vol.35, No.1, th

pp.792-795.

United States Environmental Protection Agency, 1997. Metal-enhanced dechlorination of volatile organic

compounds using an aboveground reactor: Innovative technology evaluation report, Office of Research

and Development, Washington, D.C., EPA/540/R-96/503.

Discussion

Cores collected from two commercial sites showed little change in bacterial activity over the upgradient

interface; however, no information was collected for the downgradient side of the treatment wall. 

Although there were numerous monitoring wells situated throughout the Borden site, there was no

noticeable affect on the hydrologic flow regime. Most are small-diameter shallow wells that were installed

for another purpose prior to constructing the wall.

Treatment walls can be used in low-permeability or fractured-bedrock aquifers if there is a measurable

gradient. Fractured bedrock is more difficult, although Gillham mentioned a treatment wall currently being

installed at a site with fractured bedrock. The fractures are being sealed by jet grouting, and reactive

treatment materials are being jetted into the fractures. Hydrofracturing can be used to widen fractures before

injecting materials.

There is a sharp pH gradient within the iron wall. The pH-neutral groundwater typically increased to pH 9-

10 within the first 10-15 cm of the upgradient side of the wall.
Treatment Walls NATO/CCMS Pilot Study Phase III

 6 Rue De Wattford, F92000 Nanterre, France, tel: 33/14-7764-262, fax: 33/14-9069-734, e-mail: gerard.evers@ 1

soletanche-bachy.com

22

Practical Solutions for the Treatment of Polluted Groundwater

Gérard Evers1

The remediation of polluted sites and its treatment constitute a complex problem, and every case is special.

Many solutions are proposed to clean or to isolate pollutants, such as biological treatment, soil washing or

flushing, electrochemical treatment, vacuum extraction, stabilization, or containment.

Matching a site and a remediation technique depends on many parameters. Design studies for the

improvement of these sites take into account not only the chemical context, but also the geology and

hydrogeology as well. In many cases, it is necessary to preserve the groundwater flow. 

Soletanche Bachy developed several innovative solutions in this field. Drainage trenches, used either alone

or in combination with cut-off walls, allow controlling and extracting pollutants while the local groundwater

regime is maintained.

Extensive research over the last ten years about the behavior of dissolved pollutants, organic and mineral,

ended up with a variety of materials able to trap pollutants in drainage conditions (ECOSOL and IRIS

materials), and in a process of underground water seepage control and treatment called the “drain panel

process.”

Trapping mechanisms

Several physical and chemical mechanisms are involved to trap pollutants in groundwater. Precipitation,

adsorption, and ionic exchange are the most important.

Precipitation

Chemical precipitation is mainly used to eliminate dissolved heavy metals, such as iron, nickel, copper, lead,

trivalent chromium, and hexavalent chromium.

These cations are often precipitated in the form of metallic hydroxides in chemical reactions controlled by

the pH. In cement-based materials, lime solubility dictates a pH of 12-12.5. However, this reserve of alkali

in cements, although large, is not infinite, and alkaline buffers may be incorporated in trapping materials

to extend their precipitation ability with time. In some cases, specific reagents are employed to react with

pollutants and to precipitate into a insoluble mineral forms.

Adsorption

Adsorption is a physical mechanism based on the properties of some porous materials to fix molecules on

their surface. Specific surface governs this mechanism. Attractive forces have different origins:

 • physical bonds by pores of similar sizes to the target molecules;

 • Van der Waals forces (electrical trapping) is applicable to polarized molecules;

 • surface affinity—hydrophobic organic molecules have affinities for sites on activated carbon.
Treatment Walls NATO/CCMS Pilot Study Phase III

 ECOSOL and DRAIN PANEL are patented. 2

23

Adsorption on activated carbon is a widely used method for removing organic pollutants dissolved in waste

water. Adsorption capacity values vary depending on the compounds to be removed. For instance,

adsorption capacity of chlorophenol is six times higher than that of butylacetate in same conditions. Some

clays are able to adsorb cations such as cadmium, strontium, mercury, nickel, and zinc.

Ion exchange

This process is employed for removing dissolved anions in polluted groundwater. Removal method can be

percolation through specific resins or clays. This mechanism is well adapted to eliminating cyanides. Two

processes are described here to implement these pollutant trapping drainage materials: the drain panel and

the drainage trench. 

The Drain Panel Process2

In order to avoid spreading an industrial or accidental pollutant, watertight confinement barriers are often

installed together with a pumping system, to keep the water table within the confined area below the outside

water table. A water treatment installation will clean the pumped water before returning it to the ambient

groundwater. But this classical concept has several inconveniences, which can be solved by the drain panel

process.

The drain panel process consists of drainage panels separated by watertight cement slurry panels. A conduit,

at the bottom of the watertight panel, establishes the hydraulic continuity between the two drainage panels

on either side. A valve, operated from the surface, can be installed to regulate this hydraulic continuity.

The drain panel can be built in combination with the classical confinement barrier, or in combination with

a partial barrier, which will then form the so-called “funnel-and-gate” solution. The drain panel system can

be used to considerable depths (30 meters or more).

Construction Method

The watertight cement slurry panel is excavated first. When the excavation is completed, a horizontal

conduit (200 to 400 mm diameter), linked to a vertical standpipe, is installed at the base of the panel. This

conduit has destructible parts closed off with caps at both ends. A temporary inflatable plug avoids any flow

through the conduit. 

When the slurry has set, the adjacent drainage panels can

be excavated using a biodegradable drilling mud. The

sealing slurry, vertically covering the destructible parts

located at the ends of the conduit, is excavated, and the

destructible parts are cut, either by the excavation tool or

by another suitable method. 

When the excavation of the drainage panel is completed,

it is filled with appropriate drainage material, and

conductivity between drainage panels is established by

withdrawal of the temporary inflatable plug (next three

figures). 
Treatment Walls NATO/CCMS Pilot Study Phase III

24

Drainage Material 

Drainage panels can be filled with trapping materials

adapted to the nature and content of organic or mineral

pollutants (figure below). 

If required, successive drainage panels with different

retention filters can be installed to clean the groundwater. In this case, appropriate positioning of the

conduits in the slurry panel may regulate the flow and

create an efficient volume for the active barrier.

Filter Cartridges

In this case, the whole treatment system, with conduits,

filter retention tanks and monitoring equipment, is

installed in the cement slurry (see figure at right).

Upstream, intermediate, and downstream piezometers

allow permanent monitoring of the pollutant concentration at each point of the treatment system. The filter

retention tanks are designed for an easy exchange of the

filter cartridges when they are saturated.
Treatment Walls NATO/CCMS Pilot Study Phase III

25

Other Applications for Drain Panels

Underground construction often requires temporary, watertight, cut-off walls around the excavation pit. But

without precautions, such cut-off walls will create a permanent perturbation in the groundwater flow, and

consequently will have a lasting influence on the hydraulic environment. 

The drain panel system will probably reestablish the

hydraulic continuity between the upstream and downstream aquifers. 

For a temporary cut-off wall, drain panels can be

installed at different points along the perimeter. Each

drain panel is composed of a central watertight cement

slurry panel and two drainage panels (one at each side of

the cut-off wall). While lowering the water in the

excavation pit, hydraulic connectivity through the drain

panel is avoided by temporarily plugging the conduit.

When construction is completed, the hydraulic

connectivity between upstream and downstream can be

reinstated by withdrawing the plug (figures at right). 

A possible solution for permanent works is shown on the

figure below. Drain panels are installed at each part of

the work, and the hydraulic continuity is reinstated, by

installing a drainage pipe at slab level and connection

with the drainage panels. 

Treatment of Chrome Pollution by Pollutant-Trapping

Drainage Trench

The purpose of the pollutant-trapping drainage trench is to

intercept polluted groundwater and treat it in situ. The

trench does this without maintenance, staff, or energy

input. The example given here after was applied to solve a

chromate-polluted groundwater problem.

Pollution Source and Type

The pollution was caused by materials containing hexavalent chrome being used in the construction of the A22

motorway in northern France, near Lille.

The toxicity of the material, which was known at the time

of construction, prompted engineered safeguards in the fill.

Unfortunately, the design of the central platform did not

completely control seepage. Soon after the road was
Treatment Walls NATO/CCMS Pilot Study Phase III

26

opened, there were indications of chrome (VI) in the seepage water at some spots.

In 1987, a company located below the motorway (downgradient of groundwater flow) reported chrome (VI)

in its drainage water, and this pollution was accompanied by bare patches in the vegetation of the motorway

bank.

Site Characteristics

The ground under the road embankment consists of 4-5 meters of sandy silt overlying Flanders clay. The

water in this silt flows gently towards a banked slope some 10 m downgradient of the fill, directly

overlooking the company car park.

Over time, rain infiltrating into the fill leached into the soil. This water, contaminated with chrome (VI),

percolated into the aquifer despite the layer of rolled shale at the base of the fill. Water in the toe ditch below

the bank had the characteristic lemon yellow coloring of chromate pollution. In dry weather, patches of

yellowish efflorescence appeared on the slope. Seeps of polluted water that killed off the vegetation could

be seen on the slope above the company car park.

The site was surveyed to explore the extent of the pollution. Auger holes were sunk for soil samples.

Chemical analysis revealed serious soil pollution, with chrome (VI) present in concentrations of up to

1,000 mg/kg, but disappearing beyond a depth of 5-6m because of the impervious Flanders clay bed.

Groundwater samples taken from piezometers yielded very high chrome (VI) contents, practically up to

2,000 mg/l. Chrome (VI) is soluble and highly toxic, and the maximum permitted level in drinking water

is 0.05 mg/l.

Engineering Design

The proposed solution consists of intercepting the

polluted groundwater with an Ecosol drainage

trench designed to trap chrome (VI). More precisely,

a trench is dug and backfilled with a pervious,

porous, pollutant-trapping material (figure at right).

The required material is obtained by mixing sorted

gravel with a special slurry from the Ecosol range.

These slurries are special formulations developed

and patented by Soletanche Bachy to suit the exact
Treatment Walls NATO/CCMS Pilot Study Phase III

27

nature of the pollutants in question: radioactive elements, heavy metals, or organic compounds.

The gravel grains coated with pollutant-trapping slurry bond together so when the mix hardens, they form

a rigid, porous structure through which water readily circulates. As the water seeps through, the chemical

pollutants are trapped on the film of hardened Ecosol formulation adhering to the grains of gravel.

Preparation of Ecosol Material

Several crucial parameters must be controlled when

preparing the Ecosol material (see figure at right):

 • Slurry/gravel weight ratio. This ratio must not be

too high, or the structure will not be sufficiently

porous. Nor must it be too low, or the gravel grains

will not be completely coated.

 • Gravel size. For any given ratio between a unit

volume of slurry and the weight of gravel, a finer

gravel size will increase the active surface area of the

structure and reduce the thickness of the hardened slurry coating the gravel grains. This speeds up the

trapping reactions. On the other hand, it reduces the permeability of the structure, with the attendant risk

of the pores becoming clogged.

 • Selection of gravel and slurry. The bond between the film of hardened slurry and the gravel skeleton

must be sound and durable. The use of cement slurry and compatible gravels ensures durability,

provided the same precautions are used as when designing standard concrete mixes.

 • Rheological properties of slurry. The rheology of the freshly-mixed slurry paste must be designed to

form a stable film coating the gravel. Therefore, apart from the specific pollutant-trapping substances,

the mix must be designed so that the initial stiffness of the paste when the gravel is incorporated is not

too great, nor too low.

 • Slurry formulation. All types of Ecosol slurry can be used, provided their rheological properties are

suitable. The exact formulation of the slurry depends on pollutant types and concentrations and required

trapping capacity. The mix design is done in the laboratory. In this instance, the cumulative quantity of

chrome (VI) trapped per cubic meter of material could be determined. The Ecosol material is prepared

by adding the various reagents in the composition at the mixer. This slurry is then used for coating the

gravel. This is done in a truck mixer, so that the sorted gravel is delivered, the slurry added, the mix

blended and the material delivered to the point of use all in one vehicle.

Trenching

On site, the trench is dug by a machine capable of

digging to a depth of six meters. The mobile frame

carries a chain and square-sided chute to guide the

backfill material from the top hopper into the trench. The

Ecosol-coated gravel is brought in by the truck mixer to

the trench hopper, and the blended material flows like

concrete. The result is a trench backfilled with Ecosolcoated gravel.
Treatment Walls NATO/CCMS Pilot Study Phase III

28

There were two zones: (A), where groundwater is

intercepted by two trenches, because this area has the

highest chrome (VI) concentrations; and (B), where the

flow is intercepted by a single trench.

The 0.25-m wide drainage trenches were widened at

the top. This extra thickness of trap potential deals with

surface water contaminated with chrome (VI) liable to

seep into the aquifer.

Trench Performance Monitoring

Three piezometer profiles perpendicular to the trench monitor the trend in chrome (VI) concentration upand downgradient of the of the trench. In total, 11 piezometers were installed, and each month, their water

levels are recorded and water samples taken for laboratory analyses.

Analysis of these data confirm the existence of an hydraulic gradient, perpendicular to the trenches, from

the embankment downwards to the car park location.

The following figure shows the chrome concentration upstream and downstream of the barrier, during the

18-month observation period following the completion of the construction. The figure shows the good

performance of the curtain: the chrome concentration remains at a high—but variable—level upstream of

the curtain, while negligible concentrations are recorded downstream.

The measurements are still going on, and they confirm these findings. The Ecosol material confers a high

trap potential, around 15 kg of chrome (VI) per linear meter of trench.
Treatment Walls NATO/CCMS Pilot Study Phase III

29

Discussion

In response to a question, Evers explained that the treatment wall was excavated using a trenching machine

capable of excavating only to a depth of 10 m. The overall problem with the permeable treatment wall is that

the reactive material is expensive (US$800/m ), but not all of it is receiving contaminated groundwater. 3

Because the soil is not uniform, there are preferential flow pathways, and approximately 10% of the wall

handles 90% of the contaminated groundwater. The advantage of the funnel-and-gate system is having

successive reactive materials that receive contaminated groundwater flow.

There are at least three options for constructing a treatment wall to optimize cost-efficiency: (1) construct

one continuous trench and fill it with a mixture of iron and sand; (2) use pea gravel screens upgradient to

average the flow rate through the wall; or (3) construct a wall of varying thickness to account for variations

in groundwater flow. As long as substantial information on groundwater flow was available, plans can be

made for treatment walls having variable thicknesses. Generally, however, funnel-and-gate technology has

an advantage in evening out flow through the wall, but may magnify uncertainties.

There was some concern that if the hydrology of a site is not understood, use of a funnel-and-gate system

can retard normal groundwater flow, causing an increase in upgradient thickness of the water. The

hydrology of a site must be modeled carefully prior to constructing a treatment wall. Modeling and the

design of a funnel-and-gate system can be very expensive, perhaps accounting for up to 20% of the overall

construction costs (compared to only 10% for slurry walls).
A B C

1.07 1.23

11.8

0.12 0.5 0.032

0

1 0

A Fe [m 2/g]

t

1/2n [h*m 2/mL]

Treatment Walls NATO/CCMS Pilot Study Phase III

 Institut für Wasserbau, Lehrstuhl für Hydraulik und Grundwasser, University of Stuttgart, Pfaffenwaldring 61, 70550 1

Stuttgart, Germany, tel: 49/711-685-4714, fax: 49/711-685-7020, e-mail: ww@iws.uni-stuttgart.de

30

Figure 2. 

Reactive Materials

Degradation of TCE at Zero-Valent Iron:

Chemical Processes Effecting the Design and Performance

of Permeable, Reactive Fe(0) Walls

Wolfgang Wüst, O. Schlicker, and A. Dahmke 1

Permeable reactive walls present a cost-effective alternative to classical pump-and-treat technologies for the

removal of dissolved contaminants from groundwater plumes, provided that a fairly passive long-term

performance can be guaranteed. Many materials are under research for use in such reactors, but zero-valent

iron has been studied most intensely, since it is commercially available at low costs, and a variety of

contaminants can be immobilized by reduction reactions (Dahmke et al., 1997). Based mainly on the results

of our own studies, we describe the degradation of chlorinated aliphatic hydrocarbons, the most important

application of Fe(0) reactors, and in this context, we discuss the chemical processes effecting the design and

performance of permeable reactive Fe(0) walls.

Degradation of chlorinated hydrocarbons with zero-valent iron is known to be a heterogeneous surface

reaction, and thus dechlorination can only take place after the chlorinated compound has been associated

at the iron surface (Matheson and Tratnyek, 1994). Kinetics is described as pseudo-first-order (k [1/h]) obs

with respect to dissolved concentrations, and k is reported to be proportional to the iron surface (BET) obs

concentration (Gillham and O ́Hannesin, 1994; Sivavec and Horney, 1995). Different commercially

available irons show a broad variety in costs as well as physical properties (Figure 1). 

Figure 1. Comparison of different irons

Iron Spec. surface area Porosity Costs Costs

source [m /g] [%] [DM/t] [DM/m ] 2 3

A 0.12 37 800 3,400

B 0.5 58 450 1,400

C 0.032 39 500 2,300

D 0.63 0.74 300 600

Testing iron from different producers, we found

considerable differences in reactivity with respect

to TCE degradation. Differences in iron surface

area concentrations could not sufficiently account

for it, since we normalized k to this parameter obs

(Figure 2). The normalized first-order rate

constants were in the same range as reported for

different iron sources in North America (Johnson

et al., 1996).There a various other parameters like

foreign elements and grain geometry that may

affect reactivity. Even iron from the same source

showed differences in reactivity.

In low mineralized waters competitive, cathodic

reactions may occur that can affect the performance of permeable, reactive Fe(0) walls. Oxygen
0 0.2 0.4 0.6 0.8 1 1.2 1.4 1.6

V Gas [mmol/d/kg]

a

b

d

G as developm ent at different com m ercial irons (a, b, d)

in colum n experim ents (O 2-saturation, low m in. water)

pH 4 / no TCE

pH 7 / no TCE

pH 7 / TCE in: 0.1m Mol

C om position of gases from iron source b

(initial pH 7, low m in. w ater) 

1E-3

1E-2

1E-1

1E+0

1E+1

1E+2

H2 Sum HC CH4 C2H6 C2H4 C3H8 C3H6 C4H10 HC>C4

Gas compound [% mol]

b (no TC E) b (TC E in: 0.1 m M ol)

C om position of gases from iron source d

(initial pH 7, low m in. w ater) 

1E-3

1E-2

1E-1

1E+0

1E+1

1E+2

H2 Sum HC CH4 C2H6 C2H4 C3H8 C3H6 C4H10 HC>C4

d (no TCE) d (TC E in : 0.1 m M ol)

Degradation kinetics of TCE for different inflowing concentrations

0

50

100

150

0 10 20 30 40

time [h]

C

in =13

0

500

1000

1500

C

 in 

=14

TCEin [uMol]: 134 1450

k

ol [uMol/h]: 70.5+/-43.3 59.1+/-5.4 

K1/2 [uMol]: 761+/-152 893+/-251

k

1l [1/h]: 0.093 0.066

Treatment Walls NATO/CCMS Pilot Study Phase III

31

Figure 3. 

Figure 4. 

Figure 5. 

Figure 6. 

consumption was found to be fast in our

column experiments, leading to a zone of

Fe(III) precipitates at the column inlet,

where hydraulic conductivity was reduced

considerably (Dahmke et al. 1997). Effects

on the degradation kinetics in Fe(0) reactors

were not detected, since only a few

centimeters are affected by aerobic

corrosion. Formation of hydrogen by

anaerobic corrosion is thermodynamically

favorable, but it depends strongly on the

voltage differential. For different irons, we

found dependence of reactivity on the iron

source, and pH at the inflow ranging from

no anaerobic corrosion to a strong gas

production (Figure 3), which affects

hydraulic conductivity and discharge of

gaseous TCE from the column. Hydrogen is

the dominating compound in the gas phase,

and low-weight aliphatic hydrocarbons (total

HC) consisting exclusively of saturated HC

account for less than 1 % mol gas, fairly

independent of the iron source (Figure 4,

Figure 5). TCE, at a concentration of 1

mMol in the column feed, increases the

molar fraction of the total HC to 1-3 %, and

short-chain unsaturated compounds (ethene,

propene) appear in the HC pattern,

accounting for TCE degradation. Further

column experiments discussed below were

performed with iron from a source that was

found to have the highest voltage differential

with respect to anaerobic corrosion, and a

fairly high normalized k . obs

Looking at the dechlorination kinetics in more detail,

we found that the degradation rate was levelling off

at a higher concentration due to saturation of reactive

sites at the iron surface. An enhanced model

accounting for first-order sorption, desorption, and

reaction of sorbed TCE (Figure 6, Figure 7) was

fitted to the data with good success by the

AQUASIM model. Parameters for zeroth- and firstorder kinetics with respect to dissolved TCEconcentration are estimated. Since the dominant

chlorinated intermediate, cis-DCE, did not exceed

1% of the initial TCE concentration, first-order

build-up and degradation was found to be sufficient

to fit the measured data. Neither the intermediate cis-DCE nor the kinetics of parent TCE were significantly

affected by a 5-fold increase in flow velocity (Figure 8). Cis-DCE as parent compound was degraded faster

than TCE both in the zeroth-order and first-order region (Figure 9) in the column experiments. This is the
TCE and nitrate reduction at commercial iron

0.0

0.2

0.4

0.6

0.8

0 10 20 30 40

time [h]

C

3 - , N

4

+ [mM

6

8

10

12

14

NO3 NH4 TCE pH

Treatment Walls NATO/CCMS Pilot Study Phase III

32

Figure 7.

Figure 8. 

Figure 9. 

Figure 10. 

reverse of the kinetics reported for these compounds

(Johnson et al., 1996). In the binary component

experiment, degradation rates of both compounds were

reduced and competition between TCE and cis-DCE

will be discussed in more detail (Wüst et al., in

preparation). Since groundwater plumes contaminated

by chlorinated hydrocarbons usually contain a variety of compounds, competition between these single

compounds has to be considered.

We were also testing the effect of different groundwater constituents (chloride, phosphate, sulfate, chromate,

nitrate) on the degradation of TCE in column experiments, where the dissolved concentration of the

respective groundwater constituent is gradually increased. 

Abiotic nitrate (Figure 10) and chromate (data not

shown) reduction at zero-valent iron were found to

be faster than the degradation of TCE. Nitrite as an

intermediate could be detected only in traces, and all

nitrate-N was reduced to the terminal N(-III).

Reduction of 50 mg/l nitrate (German drinking water

standard) can result in 13.3 mg/l ammonium entering

the aquifer. Since pH is driven up to values greater

than 10, ammonia is likely to form. High inflowing

nitrate concentration leads to inhibition of both the

reduction of nitrate and TCE by zero-valent iron.

Consequently, the TCE concentration front is

passing through the iron column ahead of the nitrate front (Figure 11). Passivation is due to the formation

of a considerable amount of iron precipitates (dissolved iron <0.01 mg/l). They are possibly more oxic due

to the higher redox potential both in solution and at the iron surface, and thus less conductive.
Effect of nitrate reduction on the degradation of TCE

(PV: pore volumes permeated at 50 mg/L nitrate) 

-3

-2

-1

0

ln(TCE/TCEin)

-3

-2

-1

0

0 5 10 15 20 25

time [h]

(

3 - / N

3 - in )

TCE (no NO3

- )

Effect of inflowing phosphate poncentration on the 

degradation of TCE at Fe

0.01

0.10

1.00

10.00

100.00

0 1 10 100

PO4 in [mg/L]

K ol [uMol]

K 1l [1/h]

Effect of SO4 at various inflowing concentrations on the 

degradation of TCE at Fe in column experiments

0

2

4

6

8

10

12

14

16

18

20

0 5 10 15 20 25 30 35 40

time [h]

TCE [mg/l]

1000 mg/L SO4

500 mg/L SO4

no SO4 (post)

100 mg/L SO4

no SO4 (ante)

0

1

2

3

4

5

6

7

8

half life [h]

0 4 7 23 43 45

P o re v o lu m e s s in c e in filtra tio n o f 

12 00 m g/L S O 4

2 - + 60 0 m g/L HC O 3- + 1 00 m g/L NH 4 C l

Treatment Walls NATO/CCMS Pilot Study Phase III

33

Figure 11. Figure 12. 

Figure 13. Figure 14. 

Phosphate can affect the performance of iron walls only at higher inflowing concentrations (>1 mg/l

phosphate). At 10 mg/l, we observed a decrease in both, zeroth- and first-order rate constant (Figure 12).

100-mg/l phosphate in the feed further reduced zeroth-order degradation but so far did not effect first-order

degradation in the back part of the column. Phosphate is sorbing, and forms vivianite, which was detected

by XRD in addition to magnetite. Formation of these precipitates could be predicted by the geochemical

equilibrium program PHREEQEC. In contrast to phosphate, loss of sulfate from solution could not be

observed. Considerable decrease in the degradation of TCE, however, was found at a concentration of 500

mg/l sulfate (Figure 13). Higher sulfate concentration (1,000 mg/l) did not enhance the passivation of iron.

Sorption of sulfate was found to be reversible, and the column fully recovered after its desorption from the

iron surface. 

Abiotic sulfate reduction was not observed, but microbiological transformation to sulfide took place in one

of our columns to some extent. There we found degradation rates to increase (Figure 14), which we attribute

either to the enhancement of abiotic degradation because of sulfide formation (high electric conductivity)

or to co-metabolic TCE degradation. Enhancement of reactivity was also attained by palladizing zero-valent

iron. First-order kinetics was stable over a wide range of flow velocities and no maximum degradation rate

(zeroth-order kinetics) could be detected even at high inflowing TCE concentrations (Figure 15). Increasing

chloride concentrations confirmed the results known from aerobic anaerobic corrosion studies: the firstorder rate constant increased by 2 in the range of 0-100 mg/l chloride (Figure 16).

In a low-mineralized groundwater environment, we expect long-term stability of dechlorination rates in

permeable, reactive Fe(0) walls and hydrocarbons as final products enter the aquifer. Description of the

degradation kinetics by a pseudo-first order model was not sufficient at higher TCE concentrations, where

zeroth-order kinetics was observed. We succeeded in fitting the data by a Monod-type model. Competition
Degradation of TCE at Palladized Iron (0.01% Pd)

 k1l = 0.71+/- 0.12[ h-1 ]

Co = 28mg/l v = 57cm/h

k1l = 0.70 +/- 0.16 [h-1 ]

Co = 196mg/l v = 44cm/h

k1l = 0.75 +/- 0.06 [h-1 ]

Co = 20.5mg/l v = 11cm/h

k1l = 0.79 +/- 0.25 [h-1 ]

-10.0 Co = 9.2mg/l v = 1.88cm/h

-9.0

-8.0

-7.0

-6.0

-5.0

-4.0

-3.0

-2.0

-1.0

0.0

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15

k1l = 0.82 +/- 0.056 [h-1 ]

Co = 228mg/l v = 11.6cm/h

0 11 32 41 44 50 61 69 69 74 77 79 85

0

1

2

3

4

5

6

no Cl- 2 5 mg/l C l1 0 0 mg/l C l500 m

Treatment Walls NATO/CCMS Pilot Study Phase III

34

Figure 15. Figure 16. 

among different chlorinated compounds has to be expected as shown in the case of TCE and cis-DCE, and

has to be considered in the reactor design.

Redox-competitive inorganic groundwater constituents (oxygen, water, nitrate, chromate) can affect the

design and performance of reactive iron walls considerably. Products like hydrogen gas and ammonia can

be formed and enter the aquifer. Iron precipitates can reduce permeability and reactivity. Under certain

conditions, the concentration front of the chlorinated compounds is passing through the reactor due to

passivation of zero-valent iron. These reactions may limit the scope of permeable reactive Fe(0) walls. 

Inhibition observed by phosphate and sulfate seems to be of minor importance. At ambient groundwater

concentrations, phosphate did not effect degradation kinetics of TCE, since it was precipitated at the front

of the column. No abiotic reduction was observed for sulfate (as for nitrate and chromate), but degradation

rates were enhanced by microbiological sulfate reduction. 

We expect that microbiological reduction of nitrate and chromate in the groundwater plume can reduce the

impact of these anions on the performance of reactive iron walls. Passivated walls can possibly be

regenerated by chloride, which is known to break down passive films on zero-valent iron, and which we

found enhanced degradation rates in our experiments. We further assume that ascorbic acid will recover

degradation rates due to its reducing and complexing properties. 

References

Dahmke, A., Schlicker, O. & Wüst, W., 1997. Literaturstudie „Reaktive Wände-pH-Redoxreaktive Wände”

Berichte LfU.

Matheson, L.J. & Tratnyek, P.G., 1994. Environ. Sci. Technol. 28, 2045-2053.

Gillham, R.W. & O ́Hannesin, S.F., 1994. Ground Water, 32, 958-967.

Sivavec, T.M. & Horney, D.P., 1995, Prepr. Ext. Abs. ACS Special Symposium, Anaheim, 695-698.

Johnson, T.L., Scherer, M.M. & Tratnyek, P.G., 1996. Environ. Sci. Technol. 30, 2634-2640.

Wüst W., Schlicker, O. & Dahmke A. in prep. Kinetics of the Stepwise Degradation of TCE and cis-DCE

at Commercial Iron in Batch and Column Experiments.

Dahmke, A., Schlicker, O. & Wüst, W., 1997. In: Grundwassersanierung 1997, IWS SR 28, 324-341.
Treatment Walls NATO/CCMS Pilot Study Phase III

35

Schlicker, O., Wüst, W. & Dahmke, A., 1998. TerraTech, 01, 43-46.

Discussion

Paul Bardos asked for an explanation of Figures 3, 4, and 5, which show a system with just oxygenated

water and iron producing trace amounts of short-chained hydrocarbons. Wüst said that the iron contained

3 percent carbon, which may have been the source of the hydrocarbons. Alternatively, there may have been

organics in the gas phase that were reduced.

Hermann Schad also asked for an explanation of the apparent conversion of 100% of the nitrate to ammonia,

while column experiments using site groundwater showed that much less nitrate had been converted. Wüst

responded that the difference is probably due to microbiological reactions that reduce more of the nitrate.
Treatment Walls NATO/CCMS Pilot Study Phase III

 Department of Earth Sciences, University of Wales, Cardiff, P.O. Box 914, Cardiff, CF1 3YE, U.K., e-mail: 1

liyuan@cardiff.ac.uk

 Environmental Sciences Division, Oak Ridge National Laboratory, P.O. Box 2008, Oak Ridge TN 37831-6036 2

36

The Treatment of Groundwater with Mixed-Wastes:

Reductive Dechlorination of TCE and Reductive Precipitation of Uranium

Liyuan Liang and B. Gu 1 2

Abstract

The overall goal of this project was to perform a pilot-scale field evaluation of the reactive barrier

technology at the Bear Creek Valley, Tennessee. The groundwater plume at the site contains mixed wastes

of uranium (U), technetium, organic solvents and a high concentration of nitrates. This study was undertaken

to determine the effectiveness of zero-valent iron (FeE) and several adsorbent materials for the removal of

radionuclides, nitrates and chlorinated volatile organic compounds from the contaminated groundwater.

Several types of reductive and adsorbent materials were evaluated, including FeE filings, peat materials,

ferric oxides, and Cercona Bone-Char. Results indicate that FeE is a promising barrier material to extract TM

U from contaminated groundwater and is more effective than the adsorbents, particularly at high U

concentration. A combination of FeE and peat materials was effective in removing both nitrate and U from

the contaminated groundwater. Organic solvents, such as trichloroethylene (TCE), trichloroethane and

tetrachloroethylene, were all effectively degraded in the contaminated groundwater with FeE.

 For uranium, almost 100% was removed through reactions with FeE at an initial concentration of up to 84

mM (or 20,000 mg U/l). Results from the adsorption and desorption kinetic studies and spectroscopic

studies demonstrate that a reductive precipitation of U with FeE is the major reaction mechanism. Only a

small percentage (<4%) of uranyl appeared to be adsorbed on the corrosion products of FeE. Experiments

showed that the reduced U(IV) species on FeE surfaces could be reoxidized and potentially remobilized if

re-exposed to atmospheric oxygen. However, the remobilization may not occur if the U is sequestered by

co-precipitation with ferric oxyhydroxides. Results of this study demonstrate that FeE can be an effective

medium for U removal from mixed waste in passive groundwater treatment systems.

INTRODUCTION

In situ permeable reactive barrier technology (based primarily on zero-valence iron, FeE) has been identified

as a potentially cost-effective, passive treatment technology for contaminated groundwater [1-2]. Interest

in the technology by both private industries and federal agencies has generated extensive research activities

in the last few years. Initial laboratory research has been carried out to determine the rates of the reactions

for both the destruction of chlorinated solvents, such as trichloroethylene (TCE) and the immobilization of

metals (e.g., Cr(VI), Tc(VII), and U(IV)) by FeE [3-7]. The understanding of the kinetics of contaminant

removal is necessary for designing the residence time and sizing the reactive barrier.

 Following the kinetic work, mechanistic studies have mushroomed in an attempt to understand the

pathways for dehalogenation and to identify daughter products [8-9]. Although detailed reaction pathways

and mechanisms have not been determined unequivocally, the reaction is believed to be a heterogeneous

surface reaction [7-10]. Reduction of halogenated compounds is known via hydrogenolysis, in which a

halogen ion is exchanged for a hydrogen ion, consuming 2 electrons [11]. From the known partially

dehalogenated daughter products (such as dichloroethylene (DCE) and vinyl chloride (VC) from TCE),

laboratory studies show that the stepwise dehalogenation via hydrogenolysis indeed takes place on the metal

surface [10, 12-13]. Further study has indicated that a reaction pathway via reductive beta-elimination
Treatment Walls NATO/CCMS Pilot Study Phase III

37

occurs with FeE, producing acetylene [14-15]. With a series of chlorinated ethene or methane, results show

that the less chlorinated compounds, such as DCE and VC, are more difficult to dehalogenate than are the

more highly chlorinated compounds [7]. Alternative bimetallic systems (e.g., plating of small amounts of

Ni, Cu, Zn, and Pd onto FeE) have been found to accelerate the dehalogenation rates for various volatile

organic compounds. In some cases, the bimetallic systems dehalogenate the compounds that are ineffective

with FeE alone [16-18]. 

While numerous studies have been conducted in the laboratory and field to determine the degradation

mechanisms and kinetics for organic contaminants by FeE, relatively few studies have examined the

potential for using FeE to remove radionuclides and heavy metals from contaminated groundwater. It is

believed that reductive precipitation is a major pathway for the removal of Cr(VI) and Tc(VII) [5, 19-20].

However, for uranium (U) it is still unclear whether the removal of uranyl results from reductive

precipitation or from adsorption onto the corrosion products of FeE [4, 21-23]. Knowing which of these two

mechanisms dominates the reaction is necessary because they largely determine the mobility and fate of

these contaminants in a given geochemical environment. For example, if uranyl is largely adsorbed onto the

hydrous iron oxides or the corrosion products of FeE, the adsorbed U would be readily desorbed by the

presence of competing ions and complexing agents such as CO 3 and dissolved organic matter in 2-

groundwater. Furthermore, a change in the groundwater pH would profoundly influence the adsorption and

desorption behavior of U on ferric oxide surfaces. Alternatively, if U is primarily removed by reductive

precipitation, the precipitated U would not be remobilized unless the groundwater redox conditions are

changed and the reduced U(IV) species are reoxidized.

At Bear Creek Valley, Tennessee, the groundwater is contaminated with uranium, technetium, organic

solvents (such as tetrachloroethylene, TCE, and trichloroehtane) and a high concentration of nitrates. The

overall goal of this project was to perform a pilot-scale field evaluation for treating the mixed contaminant

plume with the reactive barrier technology at the site. 

The objectives of this study were to determine the effectiveness of zero-valent iron and several adsorbent

materials in removing radionuclides, nitrates and chlorinated volatile organic compounds from the

contaminated groundwater. Results for organic solvent removal have been discussed elsewhere [33], and

in this paper, we report on the reaction kinetics and mechanisms for the reduction or removal of uranyl [as

UO (NO ) ] with various iron filings and adsorbent materials. The work was part of the remedial actions 2 32

using the permeable reactive barrier technology to intercept and treat U, nitrate, and other contaminants

migrating to the tributaries of Bear Creek at the U.S. Department of Energy’s Y-12 Plant located in Oak

Ridge, Tennessee. In particular, there is a desire to remove U from the groundwater because it is a major

contributor to risk down-gradient of the Bear Creek.

EVALUATION OF REACTIVE MEDIA FOR U REMOVAL

Uranium Reaction Kinetics

The reaction rates for uranyl removal by FeE filings were studied in batch experiments by mixing 2 g of iron

filings with 10-ml U solution as [UO (NO ) ] at an initial concentration of 4.2 mM (1,000 mg U/l). Four 2 32

sources of iron filings were evaluated: (1) Master-Builder iron filings (0.4–1 mm) (Master-Builders, OH),

(2) Cercona cast iron foam (Cercona of America, OH), (3) Peerless iron filings (0.5–1 mm) (Peerless Metal

Powers and Abrasives, MI), and (4) coarse Peerless iron filings (3–12 mm). Aqueous samples were

withdrawn, filtered and U activity was then analyzed by a Liquid Scintillation Analyzer [24]. 
Treatment Walls NATO/CCMS Pilot Study Phase III

38

Figure 1. Reaction kinetics between uranyl and zero-valent

iron filings or iron oxide powder in aqueous solution. 

Figure 2. Uranyl adsorption by adsorbent materials

(Wards peat, ferric iron oxide, and Cercona bone Char) TM

All four types of FeE filings studied were

effective in removing uranyl from the aqueous

solution (Figure 1). Over 97% U were removed

within 30 min by all types of FeE filings and no

detectable amounts of U were left in the solution

phase after about 1 hour of reaction. However,

ferric oxide was not as effective as iron filings in

removing uranyl in solution; only about 15% of

the added uranyl was removed by the ferric

oxide. The reaction rate for uranyl removal by

ferric oxide appeared to be fast, reaching equilibrium in less than 1 min (only one data point

could be collected). 

These observations indicate that U was removed

by ferric oxide largely through surface adsorption, which often exhibits fast reaction kinetics.

In contrast, the reductive precipitation process usually takes longer because it involves the corrosion of FeE

and an electron transfer process from FeE to uranyl. Because the U solution was not deoxygenated before

the experiment, a time-lag would be required to establish a low redox potential. The partial U removal by

the ferric oxide could be attributed to the limited availability of surface sites; the adsorption process reached

its maximum when an excess amount of uranyl was added in the solution. Although the FeE filings have a

relatively low specific surface area (about 1 m /g) in comparison with ferric oxide powder (10.1 m /g), the 2 2

filings removed almost 100% of uranyl in the aqueous solution after approximately 30 min. The relatively

slow reaction rates and the relatively high U removal efficiency by FeE compared to ferric oxide suggest that

reductive precipitation of uranyl is the dominant mechanism for U removal by FeE. 

U Removal Efficiency and Capacity

Because FeE filings corrode in water, uranyl is expected to adsorb onto FeE-corrosion products such as ferric

oxyhydroxides. To evaluate the partitioning of U in FeE and its corrosion products, the following experiments were performed. A 10-ml uranyl solution (42 mM or 10,000 mg/l) was equilibrated with 2 g of FeE

for approximately 3 weeks on a shaker. The solid and solution mixture was then removed from the shaker,

and the supernatant was immediately decanted. The iron filings were thus separated from the corrosion

products (i.e., particulates) in the supernatant. The amount of U associated with the filterable particles was

then estimated by analyzing U concentrations

in solution before and after samples were

filtered through a 5 μm filter. 

The U removal efficiency and capacity through

the reductive precipitation and adsorption by

FeE and several adsorbent materials (ferric

oxide, peat, and Cercona bone-char) were TM

further evaluated in batch studies. Figure 2

shows the typical adsorption isotherms of

uranyl on these adsorbent materials. The initial

slope and the plateau of the adsorption

isotherms define the adsorption affinity (or

efficiency) and capacity [25]. It is evident that

all these adsorbent materials were effective at

removing U at a relatively low solution
Treatment Walls NATO/CCMS Pilot Study Phase III

39

Figure 3. U removal by various zero-valent iron filings: MasterBuilder (MB) with headspace, MB without headspace,

 and Peerless coarse iron filings

concentration (<50 mg U/l). The initial slopes (i.e., the initial partitioning coefficients) ranged from about

40 to > 10,000 ml/g, with the Cercona Fe-bone-char being the most effective at a low initial concentration. TM

However, the Wards peat exhibited an adsorption capacity that was 2-3 times higher than that of Cercona TM

Fe-bone-char and about 5-6 times higher than that of the ferric oxides on a weight basis. As the solution U

concentration increased, a relatively large percentage of U was left in the solution because of the limited

sorption capacity of the solid materials.

 In contrast, reactions between uranyl

and FeE resulted in almost 100%

removal of U in solution as shown in

Figure 3. Regardless of the initial U

concentration in solution (up to 84

mM or 20,000 mg/l), no detectable

amounts of U were found in solution

after reaction with FeE. Note that

Figure 2 is plotted against the

equilibrium concentration of uranyl in

solution, whereas Figure 3 is plotted

against the initial U concentration

added because no detectable amounts

of U could be found in the equilibrium

solution. Analysis by the phosphorescence-lifetime instrument gave not

only a high sensitivity for detecting

uranyl (on the order of parts per

trillion) but also the valence states of

U because U(IV) does not phosphoresce [26]. In order to determine if the reduced U(IV) species were

present in the equilibrium solution phase, some selected aqueous samples were also exposed to the air and

treated with peroxide (H O ) and HNO and re-analyzed by the phosphorescence-lifetime instrument. Again, 2 2 3

no detectable amount of U was found in the solution after the treatment. These results are therefore

indicative of a reductive precipitation of U by FeE rather than an adsorption process because adsorption

would have resulted in a partitioning of U in the solution phase. No maximum loading capacity may be

defined for U removal as long as sufficient amounts of FeE are present in the system to maintain a favorable

reducing environment. The adsorption process may dominate only when FeE is consumed and the corrosion

products (Fe-oxyhydroxides) are formed in the system.

We further evaluated the percentage distribution of U in the FeE and its corrosion products in suspension

because, if U was largely adsorbed on the corrosion products of FeE, the adsorbed U may be desorbed and/or

co-transported with the suspended colloidal particles. Results (Table 1) indicated that only a small

percentage of U was associated with these suspended particles (or the corrosion products) after shaking for

~3 weeks in the batch experiments. The majority of added uranyl was precipitated on the FeE. Additionally,

a desorption experiment with 0.1M Na2 3 CO indicated that U associated with suspended particles (i.e., the

corrosion products) was readily desorbed (>64%). However, only small amounts of U (<0.21%) in FeE could

be washed out with the carbonate solution (Table 1). It is known that CO3 complexes with uranyl to form 2-

negatively charged U species such as UO2 32 (CO ) , which does not adsorb on the negatively charged ferric 2-

oxyhydroxide surfaces at the given pH conditions (pH > 10) [27-28]. The zero point of charge of common

ferric oxyhydroxides is in the range of 6-8.5 [29]. Therefore, these results suggest that uranyl, which was

in the initial solution, was converted to less soluble U(IV) species by the FeE. The large percentage of

extracted U from the suspended particulates implies that U was primarily adsorbed in its VI-oxidation state

by the corrosion products, although these corrosion products only constituted a small percentage of the FeE

mass. These observations are consistent with that of Grambow et al. [21], who found that a large percentage
Treatment Walls NATO/CCMS Pilot Study Phase III

40

Figure 4. Fluorescence spectra of uranyl before and after

reactions with ferric iron oxide (hematite) or zero-valent

iron filings in solution

of uranyl was adsorbed on the corrosion products of FeE and that uranyl was only partially reduced to U(IV)

species. It is important to note, however, that we obtained these suspended particles of iron corrosion

products by vigorously shaking the FeE solution and then immediately decanting the suspension. Under

static column flow-through conditions, no suspended particles could be observed by means of turbidity

measurements; this suggests that iron corrosion products may be cemented within the FeE mass or

precipitated downgradient in the column. This cementation of iron corrosion products is desirable for

sequestering U or other contaminants in in situ groundwater treatment systems. 

Table 1. Uranium partitioning in the FeE and the suspended particles (i.e., the corrosion products

of FeE) before and after washing with 0.1M Na CO 2 3

Peerless Iron Filings Master-Builder Iron Filings

Before wash After wash U desorbed Before wash After wash U desorbed

(mg) (mg) (%) (mg) (mg) (%)

U in FeE 96.09 95.89 0.21 97.23 97.16 0.07

U in suspended 3.91 1.39 64.5 2.77 0.18 93.6

particles 

Identification of Reaction Products by Spectroscopic Studies

Reductive precipitation of uranyl to U(IV) by FeE is thermodynamically favorable according to the

following reactions:

(1) Fe + 2e ] Fe(0) EE = -0.440 V 2+ -

(2) UO + 4H + 2e ] U(IV) + 2H O E = -0.07 V at pH 8. 2 2

2+ + -

The reduced U(IV) readily forms oxyhydroxide precipitates in solution [30-31] or

precipitates on FeE surfaces. The fluorescence

spectroscopic technique was employed in an

attempt to identify the valence state of U on

FeE surfaces and in solution. The analytical

technique was also employed to evaluate the

possible re-oxidation processes of the reduced

U(IV) on FeE surfaces. It is known that only

the oxidized U(VI) gives strong fluorescence

whereas the reduced U(IV) does not. In Figure

4, the fluorescence spectra are plotted for (1) a

uranyl aqueous solution, (2) a uranyl suspension containing ferric oxide powder (hematite),

(3) a uranyl solution in the presence of FeE,

and (4) a background aqueous solution without

uranyl. The uranyl aqueous solution itself

showed strong intensity in fluorescence. In the

presence of hematite, the intensity was decreased, but no fluorescence spectra were observed after uranyl

reacted with FeE. These results are consistent with the batch kinetic and equilibrium studies, which showed

that uranyl was reduced to U(IV) by FeE. However, because uranyl is only adsorbed onto ferric oxide and

not reduced, a strong fluorescence spectra was observed in the ferric oxide systems as a result of the

adsorbed uranyl species and the uranyl remaining in solution. 

Several lines of evidence presented above have demonstrated that, in the presence of FeE, uranyl was

primarily reduced to U(IV), which was precipitated on the iron surface. This work has addressed how
Treatment Walls NATO/CCMS Pilot Study Phase III

41

Figure 5. Fluorescence spectra of U on zero-valent iron

surfaces, showing the adsorbed uranyl or the reoxidation

of precipitated U(IV) to U(VI) species

persistent these reduced U species are on the

iron surface, especially if they are exposed to

atmospheric oxygen. Figure 5 shows the

fluorescence spectra of a few solid FeE

samples at different exposure to air after

reacting with uranyl solutions. A strong

fluorescence spectrum was observed after

uranyl was reacted with a rusted FeE and then

dried in the air for 2 days. By comparison, a

much weaker fluorescence spectrum was

observed when uranyl reacted with FeE and

was then dried in a Speed-Vac for 4 h. No

fluorescence signal could be identified for the

freshly prepared U- FeE specimen. These

results demonstrate that the reduced U(IV)

species on FeE surfaces could be re-oxidized.

Furthermore, the re-oxidation rate appeared to

be relatively slow (on the order of days) in comparison with its reduction process (on the order of minutes,

Figure 1). 

Implications for groundwater remediation

Both batch experiments and spectroscopic studies showed that FeE is effective in removing U from water

under reducing conditions. The major reaction pathway is via reduction of uranyl by FeE to form insoluble

U(IV) surface species. Adsorption of uranyl by corrosion products accounts for a small percentage of total

uranyl removal. The overall removal rates are fairly fast, and the half-life of the reaction is on the order of

a few minutes, assuming a pseudo-first order kinetic reaction. Experiments also show that re-oxidation can

occur when reduced U(IV) is exposed to atmospheric oxygen. 

These results imply that using iron in a permeable reactive barrier to remove U is feasible for groundwater

remediation. However, U retained by FeE may be reoxidized and remobilized when the precipitated U(IV)

on FeE is exposed to atmospheric oxygen. The effect of dissolved oxygen in water on the rate of re-oxidation

has not been determined in this study, but will be included in future work. As long as a reducing condition

is maintained in the permeable reactive barrier by FeE, it is likely that the reduced U(IV) is coprecipitated

and eventually cemented with the FeE corrosion products. Both laboratory and field-scale experiments are

underway to determine the removal efficiency in situ, and to gauge the geochemical influence to such

technology. The long-term performance of FeE reactive barriers with respect to its efficiency, byproduct

formation, and clogging is still a matter of debate [32]. Results of this work have demonstrated that FeE is

an effective medium that can be used to remove certain redox-sensitive radionuclides and metals, in addition

to its ability to degrade many chlorinated organic compounds.

ACKNOWLEDGMENTS

Cercona Bone-Char materials were supplied by R. Helferich at Cercona of America, Inc. Technical TM

assistance by Dr. M. Dickey, X. Yin and assistance on SEM and XRD analyses by Y. Roh and S. Y. Lee are

gratefully acknowledged. We thank D. Watson for his helpful comments on the manuscript. This work was

supported by the Subsurface Contaminants Focus Area of Office of Environmental Management, U.S.

Department of Energy. Travel support for L. Liang through the UK Environment Agency is gratefully

acknowledged.
Treatment Walls NATO/CCMS Pilot Study Phase III

42

REFERENCES

1. R.W. Gillham, U.S. Patent 5266213, Nov 30, 1993.

2. S.H. Shoemaker, J.F. Greiner, R.W. Gillham, in “Assessment of Barrier Containment Technologies”,

R. R. Rumer, J.K. Mitchell (eds.). (1995) Section 11, 301.

3. R.W. Gillham, S. F. O’Hannesin, Ground Water 32(6) (1994) 958.

4. K.J. Cantrell, D.I. Kaplan, T.W. Wietsma, J. Hazardous Materials 42 (1995) 201.

5. R.M. Powell, R.W. Puls, S.K. Hightower, D.A. Sabatini, Environ. Sci. Technol. 29 (8) (1995) 1913.

6. L. Liang, J. D. Goodlaxson, in ‘Emerging Technologies in Hazardous Waste Management VII, extended

Abstracts for the special symposium, Atlanta, GA’ (1995a) 46.

7. T.L. Johnson, M. M. Scherer, P.G. Tratnyek. Environ. Sci. Technol. 30(8) (1996) 2634.

8. American Chemical Society Extended Abstracts of 209th National Meeting, Division of Environmental

Chemistry, Anaheim, CA 35 (1) (1995).

9. Emerging Technologies in Hazardous Waste Management VII, extended Abstracts for the special

symposium, Atlanta, GA (1995).

10. L.J. Matheson, P. G. Tratnyek, Environm. Sci. Technol 28(12) (1994) 2045. 

11. T.M. Vogel, C.S. Criddle, P.L. McCarty, Environ. Sci. Technol. 21(8) (1987) 722.

12. T.M. Sivavec, D.P. Horney, in Extended Abstracts of 209th National Meeting. Anaheim, CA. Division

of Environmental Chemistry. 35(1) (1995) 695. 

13. L. Liang, J. D. Goodlaxson, N. E. Korte, J. L. Clausen, D. T. Davenport, in Extended Abstracts of 209th

National Meeting. Anaheim, CA. Division of Environmental Chemistry. 35(1) (1995b) 728.

14. T.J. Campbell, D.R. Burris, A.L. Roberts, J.R. Wells. Environ. Tox. Chem. 16(4) (1997) 625.

15. A.L. Roberts, L.A. Totten, W.A. Arnold, D.R. Burris, T.J. Campbell, Environ. Sci. Technol. 30(8)

(1996) 2654. 

16. C. Grittini, M.Malcomson, Q. Fernando, N. E. Korte, Environ. Sci. Technol. 29(11) (1995) 2898.

17. L. Liang, N.E. Korte, J.D. Goodlaxson, J. Clausen, Q. Fernando, R. Muftikian, Ground Water

Monitoring and Remediation, Winter (1997a) 122.

18. T. M. Sivavec, P.D. Mackenzie, D.P. Horney, S.S. Baghel, Preprint from the 1997 International

Containment Technology Conference, St. Petersburg, FL., Feb 9-12, 1997.

19. D.W. Blowes, C.J. Ptacek (1994). ‘System for treating contaminated groundwater,’ U.S. Patent No.

5,362,394.

20. L. Liang, B. Gu, X. Yin. Separations Technol. 6 (1996) 111-122.

21. Grambow, B.; Smailos, E.; Greckeis, H.; Muller, R.; Hentschel, H. Radiochim. Acta 1996, 74, 149-154.

22. Wersin, P.; Hochella Jr., M. F.; Persson, P.; Redden, G.; Leckie, J. O.; Harris, D. W. Geochim.

Cosmochim. Acta 1994, 58, 2829-2843.

23. Bostick, W. D.; Jarabek, R. J.; Fiedor, J. N.; Farrell, J.; Helferich, R. Proceedings of the 1997

International Containment Technology Conference and Exhibition. St. Petersburg, FL 1997, 767-

773.

24. Gu, B.; Dowlen, K. E.; Liang, L.; Clausen, J. L. Sep. Technol. 1996, 6, 123-132. 

25. Gu, B.; Schmitt, J.; Chen, Z.; Liang, L.; McCarthy, J. F. Environ. Sci. Technol. 1994, 28, 38-46.

26. Dai, S.; Metcalf, D. H.; Del Cul, G. D.; Toth, L. M. Inorg. Chem. 1996, 35, 7786-7790. 

27. Hsi, C. K. D.; Langmuir, D. Geochim. Cosmochim. Acta 1985, 49, 1931-1941.

28. Ho, C. H.; Miller, N. H. J. Coll. Interf. Sci. 1986, 110, 165. 

29. Sposito, G. The surface chemistry of soils; Oxford University Press, New York 1984. 

30. Baes, C. F.; Mesmer, R. E. The Hydrolysis of Cations; John Wiley & Sons: New York, 1976. 

31. Dai, S.; Toth, L. M.; Del Cul, G. D.; Metcalf, D. H. J. Phys. Chem. 1996, 100, 220. 

32. Liang, L.; Jacobs, G. K.; Gu, B. Coll. Surf. 1997, (in review).

33. B. Gu, L. Liang, M.J. Dickey, X. Yin and P. Cameron. Preprint of extended abstract for RTDF

permeable reactive barriers Action Tam Meeting, Virginia Beach, VA. Sept. 18-19, 1997.
Treatment Walls NATO/CCMS Pilot Study Phase III

43

Discussion

Wolfgang Wüst asked whether Liang observed a maximum capacity for sorption of uranium. He noted that

he did not see a problem with leaving uranium in an iron reactor because it is very anoxic and uranium will

remain reduced unless conditions become oxidizing and mobilize the uranium. Liang indicated that leaving

uranium in the reactor is a different issue from leaving iron in situ when used for treating organic solvents

because of the environmental implications. She said that the experiments that were conducted examined the

worst-case scenario. In column flow-through experiments, no suspended solids were noted.

Timothy Vogel asked whether nitrate was transformed mainly to ammonia or whether it was converted to

nitrogen gas as well. Liang responded that the nitrate was mainly transformed to nitrite because reduction

to nitrogen takes longer. Wüst added that he did not find nitrite in his experiments. We all agree that the

reactive media determine the by-products of the reaction.
Treatment Walls NATO/CCMS Pilot Study Phase III

 TNO/MEP, P.O. Box 342, 7300 AH Apeldoorn, The Netherlands, tel: 31/55-5493-380, fax: 31/55-5493-410, e-mail: 1

h.h.m.rijnaarts@mep.tno.nl

44

Bioprocesses in Treatment Walls: Bioscreens

Huub H. M. Rijnaarts1

Abstract

The feasibility of intrinsic and enhanced bioremediation approaches for 16 contaminated sites in the

Netherlands are discussed. At least five out of 10 chlorinated solvent sites, natural attenuation can be used

as one of the tools to prevent further dispersion of the plume. At two sites stimulation of the intrinsic

dechlorination processes in a bioactive zone is required, and pilot field tests are currently under way. In three

sulphate-reducing/methanogenic aquifers contaminated with aromatic compounds, natural anaerobic

degradation of the risk-determining benzene was demonstrated not to occur spontaneously by microcosm

studies. Benzene biodegradation could be initiated by feeding small amounts of oxygen (in all samples) or

nitrate (only in one sample). Investigations at two locations contaminated with hexachlorocyclohexanes

(HCHs) indicated significant intrinsic bioattenuation of HCHs under very low field-redox conditions, and

of the HCH-degradation products chlorophenol and benzene under sulphate/iron reducing conditions.

Biostimulated zones remain required to complete the degradation of HCH and the degradation product

monochlorobenzene. 

INTRODUCTION

The ideas concerning best approaches for remediation of polluted soil and groundwater are rapidly changing.

A few years ago many engineers and regulators considered contaminated aquifers as biologically inactive

systems that could only be remediated by very drastic and expensive methods like soil vapor extraction or

large scale pump-and-treat. Nowadays, one has become aware that the natural microbial population in soil

and groundwater react actively when confronted with pollution, and thus offer new and better ways to

protect groundwater resources at acceptable costs.

Often, natural biodegradation processes can convert large amounts of contaminants without any external

stimulus. Sometimes, the natural processes can completely degrade or immobilize pollutants during their

down-gradient transport. When sufficient time and space is available, an intrinsic remediation approach can

in principal be completely protective for man and surrounding ecosystems. At other sites, the natural

biodegradation processes need to be enhanced in order to obtain a sufficient risk reduction and groundwater

protection. Such a stimulation can be performed in a bioactive zone or bioscreen. In such a zone, the

autochthonous bacteria are supplied with the appropriate electron donors, acceptors or nutrients. At some

sites, the microbial population is not yet fully adapted to the most optimal biodegradation process. In such

a case bioaugmentation of the bioscreen may be a solution. Intrinsic remediation, when necessary combined

with bioscreens, is becoming the new approach to deal with large and complex contaminated sites.

Natural and enhanced attenuation as a risk based solution for contaminated soil and aquifers is currently

being investigated at various sites in the Netherlands. A number of the 16 cases investigated by TNO will

be discussed below.
Treatment Walls NATO/CCMS Pilot Study Phase III

45

RESULTS AND DISCUSSION

Chlorinated Solvents

At the Rademarkt site (Groningen, The Netherlands) contaminated with perchloroethylene (PCE) and

trichloroethylene (TCE), mixed redox conditions control the intrinsic biodegradation processes. In a

methanogenic/sulfate reducing zone a complete reductive dechlorination reaction via vinyl chloride to

ethene and ethane occurs. However, the transformation rates of vinyl chloride observed in the field (and in

the laboratory) are too slow to prevent migration of this hazardous compound to areas to be protected. The

low amounts of DOC (< 10 mg/l) indicate a lack of sufficient amounts of electron donor naturally present.

Laboratory experiments identified that a mixture of electron-donors is most suitable to enhance the in situ

reductive dechlorination. An in situ pilot test with an anaerobic activated zone designed for complete

reductive dechlorination is planned this fall. In another flow direction in the field, the redox condition

changes from methanogenic/sulfate reducing to oxic conditions. In the reduced part, PCE and TCE are

transformed to cis-1,2-dichloroethylene (DCE) and vinyl chloride (VC), which both disappear after entering

the oxic zone. Probably, these compounds are removed by intrinsic oxidation. Hence, in this flow direction,

a complete sequential degradation of the chlorinated compounds is achieved.

At another PCE/TCE site in Maassluis, the contaminated aquifer contains high concentrations of organic

carbon (up to 700 mg/l DOC). Here, complete reductive dechlorination is observed. Most likely, this is a

result of the high amounts of intrinsic electron donors and a well-adapted autochthonous microbial population.

The results of the characterization of redox conditions and intrinsic biodegradation at 10 chlorinated solvent

sites will be presented. For at least five of these sites studied, intrinsic remediation is an important part in

an approach to effectively control the risks. At two sites at least, in situ biodegradation needs to be

stimulated in in situ activated zones to protect down stream areas. 

Intrinsic chlorinated solvent remediation is further investigated by testing electron donors and measuring

in situ hydrogen pressures in the field and in laboratory microcosms. Thus a further insight into mechanisms

involved will be obtained.

At present, new technologies are being developed by TNO for application in bioscreens and remediating

chlorinated solvent hot spots and chlorinated solvents in low-permeable soils and subsurface layers.

Combination of electro-reclamation techniques and biological methods appear to bring new solutions in the

near future.

Aromatic and Oil-Related Hydrocarbons

Aromatic compounds are often the risk-controlling compounds at oil and gas production sites, and at sites

of coating and nutrition industries. At three sites in the north part of The Netherlands, deep anaerobic

aquifers contaminated with benzene, toluene, ethylbenzene or xylenes (BTEX) have been investigated.

Under the existing sulfate-reducing conditions, the intrinsic biodegradation of toluene and ethylbenzene

could be demonstrated in the field and in microcosm studies. Benzene biodegradation could not be

evidenced with the field data. Laboratory microcosm studies with five different sediment samples were

performed. For each sample a series of comparable triplicate microcosms were incubated; in total about 300

microcosms were used. The results demonstrate thus far (200 days of incubation) that spontaneous intrinsic

anaerobic benzene biodegradation does not occur in these sediment and groundwater samples. 

Microcosm method development studies indicated that special care is required to prevent artifacts: “apparent

intrinsic anaerobic” benzene biodegradation could be demonstrated to originate from low (often not
Treatment Walls NATO/CCMS Pilot Study Phase III

46

measurable) amounts of oxygen introduced into the systems, when inappropriate materials and techniques

were used. 

Microcosms were also used to investigate possibilities to stimulate biodegradation of benzene and BTEX

compounds. Especially, addition of nitrate and low amounts of oxygen to the anaerobic samples was studied.

In one out of the five series of sediment-microcosms, nitrate-reducing benzene biodegradation appeared to

occur after lag-times greater than 100 days. In addition, anaerobic benzene degradation could be initiated

in all samples by spiking low amounts of oxygen. Further investigations are underway to elucidate

mechanisms and quantify remediation process parameters. The results are used for designing pilot

demonstration tests to be performed this spring/summer.

At an oil refinery site in the Rotterdam Harbor area, an aerobic reactive-trench bioscreen is tested for

managing a plume of the dissolved fraction of a mineral oil contamination (80% of the compounds belong

to the C6 - C12 fraction). Bench scale experiments are currently performed to establish: i) optimal grain-size

and packing density for the porous media used in the trench; and ii) optimal oxygen supply rates to

sufficiently initiate aliphatic hydrocarbon biodegradation and to minimize clogging with iron(III)oxides. 

Chlorinated Pesticides

HCH isomers are important pollutants introduced by the production of lindane (gamma HCH). At two sites,

intrinsic anaerobic biodegradation of HCH and corresponding degradation-intermediates (benzene and

monochlorobenzene) is currently being investigated. At one site, natural biodegradation processes appear

to completely degrade all compounds except the monochlorobenzene; possibly some biostimulation may

be required at the downstream end of the plume. At the other site, interception of the HCH/chlorobenzene/benzene plume is required. A bioactivated zone as an alternative to conventional large scale pumpand-treat is currently being investigated. At present, laboratory process research aimed at developing a

combination of anaerobic-microaerophilic in situ stimulation in such a bioactivated zone is being performed.

The results indicate good prospects for further development and pilot scale demonstration.

Guidelines for Application of Natural Attenuation or Bioscreens

Dutch and Nicole-supported guidelines for assessing the feasibility of natural attenuation approaches for

various contaminant situations are currently under development, making use of already existing protocols

(USA) and taking Dutch/European specific conditions, knowledge and regulatory constraints into account.

These guidelines are to become tiered decision tools. In the first steps, a minimum of data is required for

an initial feasibility screening, thus saving costs. At higher tiers, data requirements are designed to address

natural degradation under various redox-situations (including sequential redox conditions) that often occur

in European sedimentary regions.

CONCLUSION 

Intrinsic and enhanced in situ bioremediation approaches become more and more accepted as appropriate

cost-effective solutions for aquifers and soils contaminated with organic chemicals. These techniques still

need to be further developed and demonstrated and could also be expanded to control and remediate

mixtures of organic and inorganic (heavy metal) pollutants. 

Discussion

Hexachlorocyclohexanes can be degraded using zero-valent iron and without producing benzene as a byproduct. Zero-valent iron may be an appropriate reactive material for these types of compounds as long as

the mass-balance considers monochlorinated benzene.
Treatment Walls NATO/CCMS Pilot Study Phase III

47

Because there is no natural nitrate at the site, anaerobic benzene degradation was stimulated by introducing

nitrate. Even though The Netherlands has a 50 mg/l nitrate drinking water standard, the introduction of

nitrate did not pose a problem because concentrations in the stimulated system ranged from only 10-20 mg/l.
Treatment Walls NATO/CCMS Pilot Study Phase III

 This paper was prepared from a transcript of Dr. Vogel’s presentation. 1

 Rhodia Eco Services\ATE, 17 rue Périgord, 69330 Meyzieu, France; tel: 33/4-7245-0425; Fax: 33/4-7804-2430; e- 2

mail: timothy.vogel@rhone-poulenc.com.

48

Figure 1. Reactions

Reaction rates Activation energy, steric

hinderance, diffusion

limits

Compounds degraded Electrophilicity,

hydrogenation

Corrosion of material Selective reactivity,

protection, pretreatment

Metal inactivation Chemical reactivity,

redox changes —

electron sources

Figure 2. Problems and Solutions

Novel Catalyses for Reactive Barriers1

Timothy M. Vogel2

The presentation that we will make today is a combination of several different presentations, not necessarily

those planned for today. But the point is simply to talk about different types of material that might be able

to catalyze or to react with compounds of interest in barriers and in wall systems, so we can target a number

of different reactions (Figure 1), and we’ll use some information that has been collected in other laboratories

and with participants, including Nada Assaf-Anid at the University of Manhattan, Pedro Alvarez from the

University of Iowa, and Larry Nies at

Purdue University, who all have

contributed in some or the other to the

data that we will present today.

There are basically two different

concepts for the remediation of groundwater contamination: (1) Methods based

on transfer and concentration of the

contaminants based on activated carbon

—activated carbon even in a wall is not

a reactive wall but an adsorption

wall—and more innovative methods

such as mobilization of contaminants

with surfactants or in situ absorption barriers—but these methods all produce secondary waste streams that

have to be treated separately. (2) Destructive methods—such as thermal, oxidative, biological, or catalytic

dehalogenation—that are shown in the laboratory to work effectively.

What we’d like to do is to talk about these reactive systems from two points of view: what kind of reactions

we would like to observe, and what reactions we would not like to observe. The example is with the

chlorinated alkenes, where a lot of individuals do not like to see the presence of vinyl chloride. What we’ll

talk about is ways to catalyze some of these reactions and to avoid certain by-products.

Our approach is to look at the different problems

that have been associated with literature reports on

catalyzed reactions and to try to identify ways to

answer these questions. Today one of our goals is

to, in a sense, initiate some discussion, because we

are going to talk about some things that have not

necessarily been brought to a scientific, critical

review, nor applied in full-scale processes, so any

comments that you have would be gratefully

received.

We looked at reaction rates (Figure 2), and the

question often asked is, how we can do reactions

faster. In a talk presented yesterday, there was a
Treatment Walls NATO/CCMS Pilot Study Phase III

49

Figure 3. Simplified electron flow

comparison between zero-valent iron and biological reactions, where the rates were much faster with the

zero valent. And then just recently, we had another on the use of bio-screens, so there is a question about

reaction rate that has to be considered with cost—we’re not going to talk about cost—we’re going to talk

about what we can to increase reaction rates. So either we try to change the activation energy of the system,

or in the case where there is some kind of steric hindrance—where the molecules cannot come in contact

with the reactive surface—or in the same respect diffusion limits between the molecules in the bulk solution

and the site of reactivity. These are ways to address the reaction rates and maybe improve them.

Other questions are related to the compounds that are degraded. Many compounds are degraded by existing

systems, but are degraded at rates that are not acceptable, therefore some people say these compounds are

not degraded by a particular system. For example, some people say iron cannot degrade a certain chlorinated

compound, when in fact it’s more the rate that is not acceptable, not whether it can degrade or not degrade.

We’d like to separate the issue of the possibility of a degradation occurring from an unacceptable rate. Thus,

we want to examine the kinds of mechanisms and how they affect these dechlorinations or reductions. And

finally, there are issues about the lifetime of the material, and how it’s affected, and how you can engineer

support material and different catalyst systems, such that the lifetime is long enough for the application. 

In addition, we will discuss some biological aspects and using organisms coupled with other systems like

iron to deal with problems that result from the applications of catalysts.

The general schematic is a cyclic system (Figure 3)

in which an electron comes from some source, is

transferred to the molecule that is the pollutant, the

pollutant is reduced by that reaction, and the source

is then regenerated (or maybe not) (Assaf-Anid et

al.). There are two ways of looking at this: One is

based on thermodynamics, and thermodynamics

are (for those who are not aware of these systems)

the pressures pushing the system around—the

force. But then, there is also the question about the

kinetics, the rates. You can have a large force, but

if you have a small rate, it will be slow. Take a

hydraulic dam. You can have a lot of water behind

a hydraulic dam, but if the hole is small, not much water will go through it, so we need to separate

conceptually thermodynamic data from kinetic data. The thermodynamic possibility is necessary, but it

doesn’t provide you with a real process that you can apply if the kinetics are unacceptably slow.

If we look at the thermodynamics of different systems, we can see in Table 1 (taken from several literature

studies) that there are measures of different redox couples that are used to determine whether the

thermodynamics of a system is favorable not. This is the redox couple between the oxidized and reduced

phases of these compounds. Some of the data are pretty standard, for example water—the oxygen/water

couple. This provides you with some information about the potential energy—how high the water is behind

the dam—but it does not tell you anything about the rate at which these reactions will go. These are just

potential electron donors in systems, and those in Figure 4 are ones from biological systems based on a 1987

review (Vogel et al., ES&T) concerning these types of catalysts. 

Table 1. Standard potentials (EE) of biologically relevant electron donors or reductants

Reductants EE (volts)

Vitamin B12 -0.59 to -0.8

Co(I) tetraphenylporphin -0.56
Treatment Walls NATO/CCMS Pilot Study Phase III

Reductants EE (volts)

50

Ferrodoxin (reduced) -0.43

H -0.42 2

Cr(III) -0.41

NADH + H -0.32 +

Cytochrome P450 -0.30

(unactivated)

Glutathione (reduced) -0.23

Cytochrome P450 -0.17

(activated)

Fe (II) deuteroporphin IX 0.00

Ubiquinone (reduced) 0.10

Cytochrome c (+2) 0.22

Fe(II) 0.77

H2O 0.82

What is of interest is the effect that the complex has on the metal. If you talk about iron, or oxidized iron,

and there is some work with iron porphyrins, the important aspect is that the term, “iron reactive barriers,”

for example, is uninformative because it’s necessary to consider whether it’s zero valent iron; whether the

iron is complexed in some form or the other; or what the iron is doing. 

Table 2 shows redox couples between Fe and Fe in different complexes or the different ligands. Redox 3+ 2+

energy varies for the same iron couple, based upon the way that it’s complexed by the organic: the potential

for reduction or oxidation in these systems varies with the ligand. There is also some cobalt data in the table.

A metal is not in an isolated situation, but has certain complexes associated with it that may have an effect

on the thermodynamics.

Table 2. Standard reduction potential of Fe and Co to the respective bivalent ions 3+ 3+

Ligand EE Fe (+3/+2), V EE Co (+3/+2), V

1,10 phenanthroline +1.20

3 bipyridin +0.95 +0.31

water +0.77 +1.88

glycine -+0.4 +0.15

6 NH +0.37 +0.11 3

6 (CN)G +0.36 -0.80

cytochrome c +0.25

meso-tetrakis (N-methyl-pyridil) porphin +0.17

EDTA +0.13 +0.37 4-

rubredoxin -0.05
Treatment Walls NATO/CCMS Pilot Study Phase III

Ligand EE Fe (+3/+2), V EE Co (+3/+2), V

51

Figure 4. Thermodynamic potentials

iron protoporphyrin -0.12

8-hydroxyquinoline -0.15

The redox potential of your chlorinated or nonchlorinated organic that you want to reduce needs

to be examined in order to determine (based upon

your electron donor and your electron acceptor)

whether you have a potential energy production.

And again, borrowing from a review (Vogel et al.,

1987, ES&T), Figure 4 shows data where, by

classic chemical methods, you can determine the

energy that would be derived, and the thermodynamic potentials for coupling different reactions,

electron donors, and acceptors. 

For example, one reductive dechlorination is from

tetrachloroethylene (PCE) to trichloroethylene

(TCE). That is one-half of the couple, and if you

coupled it then with vitamin B , for example, 12

which is a cobalt-containing porphyrin, you would

have a favorable reaction. The distance would be

the amount of energy, the potential energy from

that reaction is about one volt. In other words, it is

a thermodynamically favorable reaction to use

vitamin B to dechlorinate PCE to TCE. This is 12

just a simple figure that has two couples, an

electronic donor and an acceptor. These types of

tables and calculations are useful because they

allow you to determine what the thermodynamic

potentials are in your system, assuming you calculate this accurately for the concentrations in your

actual system and not for standard-state

calculations that is typical in the literature. Then

you can determine in your system whether you

have a gap here, and whether you have the potential to cause the reaction that you want. Note that as you

get less chlorine on some of these chlorinated aliphatics, the redox potential decreases. 

In summary, this concept of coupling these different kinds of complexes with reductive dechlorination or

with reductions of molecules has been studied and discussed extensively in the literature. This is all in

published review articles, but it’s enough to give you an idea. Table 3 shows some of the chlorinated

compounds that have been observed to degrade and the complexes that have been reported to be responsible.

The diversity of potential catalysts for chlorinated compounds is relatively high, and from a thermodynamic

point of view, there is considerable interest.

Table 3. Reduction of halogenated aliphatic compounds by transition metal complexes

Compound Products Reductant

Methanes

Chloromethane Alkylated co-complex Co(I) chelates
Treatment Walls NATO/CCMS Pilot Study Phase III

Compound Products Reductant

52

Dichloromethane Methane Cr(II) SO

B -Co(III) 12

4

Cl-alkylated B 12

(methylcobalamine)

Trichloromethane Methane Cr(II)SO

Fe(II)P

Dichloromethane Fe(II)P

Cl-alkylated B B -Co(III) 12

4

Dichloromethane

12

Tetrachloromethane Methane Cr(II)SO

Chloroform Fe(II)P

Cl-alkylated B B -Co(III) 12

4

12

Bromomethane Methane Cr(II)SO

Alkylated co-complex Co(I)-complex 4

Dibromomethane Methane, ethene Fe(II)P

Tribromomethane Br -alkylated B B -Co(III) 2 12 12

Ethanes

Chloroethane Alkylated co-complex Co(I)-complex

1,1-Dichloroethane Ethane, ethanol Cr(II)SO4

1,1,1-Trichloroethane Ethane, ethanol, ethene, Cr(II)SO

chloroethane

1,1-Dichloroethane Fe(II)

1,1-Dichloroethane Fe(II)P

4

Hexachloroethane Tetrachloroethene Fe(II)P

Tetrachloroethene Cr(II)SO4

Bromoethane Ethane Ni(I)

1,1-Dibromoethane Ethane, ethanol Cr(II)SO4

1,2-Dibromoethane Ethene Fe(II)

Ethene Fe(II)P

Propanes

1-Chloropropane Alkylated co-complex Co(I)-complex

1,1-Dichloropropane Propane, propanol, propene Cr(II)SO4

1-Bromopropane Alkylated co-complex Co(I)-complex

1,2-Dibromo-3- Propene, allyl chloride Cr(II)SO

chloropropane 4

Note: The homolytic cleavage of cobalt-carbon bonds requires 15-30 kcal/mol and can occur at relatively low temperatures.

Figure 5 shows an example where a carbon tetrachloride (CT) is dechlorinated by cobalt in solution to

chloroform (CF) over a period of about eight hours. When that cobalt was placed into a porphyrin (Figure

6), you see that the reaction rate changes drastically and you have dechlorination within about an hour

(Figure 7 is an example of a porphyrin).
Treatment Walls NATO/CCMS Pilot Study Phase III

53

Figure 5. Carbon tetrachloride (CT) dechlorinated to

Carbon fluoride (CF).

Figure 6. Carbon tetrachloride dechlorinated to

carbon fluoride in the presence of vitamin B12

Figure 7. Porphyrin Core

Figure 8. Nitrate removal can be enhanced by

combining FeE with hydrogeonotrphic denitrifiers.

Bacteria will use cathodic H2 as electron donor to

produce harmless N instead of objectionable NH . 2 4

+

The same kind of process is observed with a large range of

metals, thus, the practical question becomes where can the

metals be placed: on what kind of surface, or in what kind of

complex in which good thermodynamics and a high kinetic rate

are possible. In other words, what we want is a lot of water

behind the dam and a large hole in the dam, so that the reaction

flows. 

Another example is with zinc (Figure 8): by itself; with carbon

tetrachloride again; and zinc with vitamin B , which goes 12

much faster. So using that concept of thermodynamics and that

aspect of kinetics, you can engineer systems where your

catalyst is potentially active from an energetic point of view,

and where the system has physical characteristics and chemical

characteristics such that the kinetics will be quite rapid.

Another proposed use of iron is with a microorganism for the

conversion of nitrates to N to avoid ammonia (Alvarez et al.). 2

Figure 8 is a schematic of the process: an organism that would

use hydrogen gas and would be responsible for this reaction.

The thought was that it would provide some synergy between

the iron and the microorganism. This synergy would allow for

some benefits in terms of the hydrogen production, in terms of

possible effects on the iron, and also possibly reducing the

competition of nitrate with other compounds like chlorinated solvents on the iron (Figure 9). In addition,

it might cause a higher reactivity for certain

compounds that don’t react very rapidly with iron,

like dichloromethane. That’s the basis for the

synergy between the two. Figure 10 shows a mass

balance in different tests: iron powder, steel wool,

iron with this organism, Pseudomonas,

Pseudomonas and steel wool, and with hydrogen.

The “Unrecovered NO ” in the figure is the lack of 3

-

the mass balance. With iron powder alone, mostly

ammonia was produced. With steel wool alone, there
Treatment Walls NATO/CCMS Pilot Study Phase III

54

• eE corrosion rapidly induces anoxic conditions.

• Bacteria may remove passivating H2 layer from FeE

surface, increasing FeE reactivity. Use of H2

increases e flow from FeE (cathodic depolarization). -

• Removal of inhibitory products by FeE enhances

bacterial participation in cleanup process (e.g., NO2

-

reduction by FeE; metals precipitation due to high pH

caused by FeE corrosion).

• Bacteria could remove pollutants that FeE cannot

(e.g., dichloromethane).

Figure 9. Basis of Synergism for FeE + Bacteria

was considerable un-reacted nitrate and ammonia.

And in the presence of the microorganism,

considerable amounts of the N2O were measured in

these mass balances. 

Figure 10. Mass balances

Discussion

In response to a question from Resat Apak, Vogel indicated that some metal chelates used as catalysts are

inside membranes. They are trapped in a system that allows the transfer of the chlorinated compound, but

keeps the solubilized porphyrins in membrane systems. Vogel indicated that there are several techniques

documented in the literature to re-oxidize the cobalt(II)EDTA to cobalt(III)EDTA, including the use of

electron donors, solid catalysts, and chemical reductants.

Wolfgang Wüst asked whether any column experiments were done before conducting the field work,

because his experience has shown that if catalysts are used, iron hydroxides or other precipitates inhibit the

diffusion (and thus reactivity) of EDTA. Schüth said that he was currently running laboratory experiments

in which several thousand pore volumes have been exchanged with groundwater. He has not observed much

precipitation in these experiments because the pH was lowered, not raised. However, microbial growth did

pose a problem for mass transfer due to diffusion.

Liyuan Liang asked whether and how much hydrogen was introduced into the system. Vogel indicated that

hydrogen had been introduced in the laboratory, and the amount was calculated by stoichiometry. In the

column and laboratory experiments, hydrogen gas was bubbled through the water to create hydrogensaturated water. However, Vogel cautioned that this approach may not work in the field.

Paul Bardos observed that many of the projects that were discussed in the treatment wall session involved

highly engineered solutions in terms of emplacing walls and improving the reactive matrix. He asked if the

advantages of these improvements were actually worth the increased costs, given that a trench with iron

filings treats groundwater reasonably well. Vogel responded that iron filings cannot treat all contaminants,

nor is there any other existing PRB suited for all contaminants and subsurface environments. Cocontaminants can cause problems, and some contaminants are less reactive than others. The contaminants

that cannot be treated by iron walls must be determined, and inexpensive reactive materials must be found

to treat them. 

Noting that the cost of iron filings for an iron wall accounts for a significant proportion of the construction

costs, Bob Siegrist asked how the use of a more expensive catalyzed metal such as palladium could be costeffective when compared to pump-and-treat methods. Vogel responded that although palladium costs more
Treatment Walls NATO/CCMS Pilot Study Phase III

55

than iron, the reactivity of palladium is also greater, so it could still be more cost effective because you do

not need as much material. Wüst noted that his tests have indicated that the cost of using 0.1 percent

palladium (the optimum amount) was approximately the same as iron, but the reactivity of the palladium

was much greater. He expressed concern, however, that the palladium could be mobilized under certain

conditions and contaminate the groundwater.

Harry Whittaker noted that there has been a lot of discussion thus far on reductive dehalogenation barriers

and asked whether oxidative dehalogenation barriers are being developed. Vogel indicated that techniques

for oxidative dehalogenation barriers have been patented recently, including systems involving Fenton’s

agent.
Treatment Walls NATO/CCMS Pilot Study Phase III

 IMES GmbH, Kocherhof 4, 88239 Wangen, Germany, tel: +497528 97130, Fax: +49752897131, e-mail: 1

hermann.schad.imes@t-online.de

 Applied Geology, University of Tobingen, Sigwartstr. 10, 72076 Tubingen, Germany, tel: +49 7071 297 5429, Fax: 2

+497071 5059, e-mail: grathwohl@uni-tuebingen.de

56

Full-Scale Projects

Funnel-and-Gate Systems for In Situ Treatment of Contaminated

Groundwater at Former Manufactured Gas Plant Sites

Hermann Schad and Peter Grathwohl 1 2

Introduction

Groundwater contamination by polycyclic and monocyclic aromatic hydrocarbons (PAH and BTEX) is

typical for many former manufactured gas plant sites. Most of the PAHs are very persistent in the subsurface

environment—they are still present in high concentrations many decades after the contamination occurred,

and they cannot be removed from the subsurface within a reasonable time period by pump-and-treat. This

persistence above all is caused by slow dissolution kinetics of the compounds from non-aqueous phase

liquids, slow diffusion of the contaminants from low permeability zones (that have accumulated the

pollutants over decades) or resistant adsorption of the contaminants by the aquifer material.

Fast remediation of such contaminations is only possible by excavating the contaminated soil. This,

however, can only be applied at shallow sites and is still expensive because of the high disposal or treatment

costs of the contaminated material. Also the location and extent of the subsurface contamination has to be

known in detail, which is not feasible at large and abandoned industrial sites. Thus, new alternative

approaches for in situ groundwater remediation have to be developed. Since the goal of remediation in

general is the protection of groundwater resources downgradient from a contaminated area, in situ treatment

may be focused on the plume rather than on the source. This can be achieved using the concept of in situ

passive treatment systems, which in general are built perpendicular to the flow direction (e.g., Teutsch et

al., 1996). The technical implementation is either in the form of a continuous permeable wall or as a funneland-gate system. In both cases no active pumping is required to move the groundwater through the treatment

zone. Within the reactive zone the pollutants are either degraded, sorbed or precipitated through biotic or

abiotic processes.

In this paper, we present an example for a funnel-and-gate system, which is currently being planned for the

long-term remediation of the former manufactured gas plant site of the city of Karlsruhe in southern

Germany.

Site Characterization

The site is located within the municipal area of the city of Karlsruhe and covers approximately 100,000 m . 2

Site investigation programs were completed in 1996 (Trischler and Partner, 1996) with the recommendation

for funnel-and-gate as the most favorable remediation technique for that site. The latter was based on a

feasibility study (IMES, 1996), in which the general and site specific advantages of passive in situ treatment

were discussed and evaluated technically and economically.

The geological and hydrogeological situation at the site may be classified as typical for the upper Rhine

valley. The aquifer has a mean thickness of about 12 m and consists of mostly sandy gravel, which is
Treatment Walls NATO/CCMS Pilot Study Phase III

57

Figure 1. Site map showing the PAH plume and the location of the funnel-and-gate system

underlain by a clay layer at a depth of 15 m below surface. Groundwater drainage is dominated by the Rhine

river northwest of the site. The mean groundwater flow direction therefore is from southeast to northwest

(Figure 1). The contamination of the site is dominated by PAHs with acenaphthene being the highest

concentrated compound of the plume extending about 400 m down gradient from the site. Several

infiltration hot spots of non-aqueous phase (NAPLs) liquids were located within the saturated zone at the

site. PAH concentrations of the soil at many places exceed 10 g/kg. Altogether approximately 55,000 m or 3 

100,000 tons of soil (within the saturated zone) are highly contaminated.

The hydraulic conductivity of the aquifer was estimated to 1.4 x 10 m/s and the hydraulic gradient is -3 

approximately 0.1%. The groundwater flow rate from the contaminated site under natural conditions is about

2.7 l/s.

Based on the environmental regulations to be applied, the groundwater contamination at the site has to be

remediated. For several remediation techniques the investment and running costs were estimated prior to

the final evaluation of the site investigation results (Table 1).

Table 1. Summary of estimated costs in DM for 50 years operation

for different remediation techniques (Trischler and Partner, 1996)

Technique Running Costs Total Costs Investment Unforeseeable

Costs Costs

Thermal Treatment 34,900,000 200,000 3,500,000 38,600,000

Containment 11,600,000 4,000,060 3,100,000 18,700,000

Pump-and-Treat 5,600,000 8,300,000 1,400,000 15,300,000

In Situ Immobilization 15,600,000 600,000 3,200,000 9,400,000

Funnel-and-Gate 7,500,000 4,100,000 1,700,000 13,300,000

With a feasibility study (IMES, 1996) it was shown, that a funnel-and-gate can be installed at the site. The

main components of such a system will be a slurry wall of approximately 240 m in length and two gate

buildings with a total volume of about 400 m . The principal questions with regard to the system design are 3

related to the decontamination process within the gates and to the hydraulics of the entire system.
Rd ' 1%Kd @

D

n

Treatment Walls NATO/CCMS Pilot Study Phase III

58

Decontamination within the gates will be based on adsorption of the PAHs on activated carbon. In order to

determine the design parameters for the gates the processes relevant for operating the gates have to be

identified and parameterized. In order to answer the questions related to the hydraulic behavior of the

funnel-and-gate a numerical model will be necessary. Based on these experimental and numerical

investigations the technical construction details have to be designed.

In the following sections the key questions related to sorption, hydraulic behavior and technical construction

will be discussed based on the results of the feasibility study.

Sorption/Retardation

The adsorption of hydrophobic organic contaminants from the aqueous phase generally increases with

decreasing solubility of the compound (or increasing octanol/water partition coefficient, Kow) and increasing

organic carbon content of the aquifer solids. Natural materials with high organic carbon content such as

coals or bituminous shales cause significant retardation of organic contaminants from groundwater. Such

materials with high sorption capacities may be used for passive removal of strongly hydrophobic

contaminants in groundwater (such as PAHs), which are the key problem at many former manufactured gas

plant sites). Much more efficient is adsorption onto activated carbon, which is already a well-established

technology for ex situ treatment of drinking water, polluted groundwater or waste water. For successful use

in a permeable wall or in a funnel-and-gate system permeability and sorptive properties of the adsorptive

wall material (adsorbent) must be optimized. Both the permeability and the sorption rates depend on the

grain size of the adsorbent, the permeability increasing with increasing grain size, and the sorption rates

decreasing with increasing grain size squared.

In general, an economically efficient in situ groundwater treatment system that relies on adsorption of the

contaminants must fulfill the following requirements:

1. Long regeneration cycles (3-10 years). Thus, high sorption capacities or adsorption combined with

biological or abiotic degradation of less sorbing compounds are required.

2. Relatively high permeability in comparison to that of the aquifer in order to prevent steep hydraulic

gradients.

3. Fast sorption kinetics in order to achieve high retardation factors even at high groundwater flow

velocities (short contact times in funnel-and-gate applications).

4. No decrease in permeability or “chemo-biofouling” of the adsorbent due to competitive adsorption of

dissolved organic matter or the growth of a biofilm that may plug adsorbent pores (although

biodegradation of contaminants would be highly desirable for a sorptive plus reactive wall).

The retardation of contaminants, such as in a sorptive permeable wall, can be calculated based on the

sorption coefficient, K :d

where D denotes the bulk density and n the porosity of the filter (in packed beds both values lie within a

narrow range). R may be interpreted as the number of pore volumes that can be displaced before break- d

through of the contaminant occurs. K denotes the ratio between the sorbed and aqueous concentrations of d

the contaminant. If sorption is nonlinear (which is often the case for strongly hydrophobic contaminants and

strong adsorbents such as granular activated carbon, GAC), K d can be calculated in a first approximation

from the commonly-used Freundlich type adsorption isotherm:
Kd ' Cs

Cw

' KFr C1/n&1

w

Treatment Walls NATO/CCMS Pilot Study Phase III

59

where C , C s w, and KFr denote the equilibrium solid and aqueous concentrations of the solute and the

Freundlich sorption coefficient, respectively. The Freundlich exponent 1/n is usually smaller than 1,

resulting in decreasing K values with increasing concentration of the contaminant. d

Activated Carbon Adsorption

Adsorption capacities for organic compounds using activated carbons are much higher than in naturally

occurring organic sorbents. Figure 2 shows literature data on adsorption of a variety of organic compounds

onto activated carbon (F300). Distribution coefficients (K ) d were calculated based on the Freundlich sorption

isotherms (K , 1/n) for an aqueous concentration of 10% of the pure compound’s water solubility. For Fr

compounds with high water solubilities (S > 1 mg/l), an almost linear inverse relationship between Kd and

S is observed (solid line in Figure 2). Since bulk density and porosity in a packed bed of granular activated

carbon are both close to 0.5, K d in a GAC passive treatment wall is approximately equal to the retardation

factor R .d

Sorption Kinetics and Permeability

The data shown in Figure 2 are valid for equilibrium sorption, which is not always applicable since the mean

residence times (contact times) in a permeable sorptive wall may be too short for complete equilibration.

Sorptive uptake of solutes by a porous adsorbent particle is diffusion limited. This results in K values that d

increase with the square root of time for short-term sorptive uptake (less than 50% of sorption equilibrium

reached). Fast equilibration would be achieved for small particles with high surface to volume ratios. Small

particles, however, result in low permeabilities. In a funnel-and-gate system the permeability of the gate has

to be as least as high as in the aquifer. In a first approximation the permeability (K) increases with the grain

radius of the adsorbent particle squared and the intergranular porosity. Since the porosity in loose packed

beds will generally lie within a relatively narrow range, between 0.4 and 0.6, the main factor influencing

the permeability is the size of the adsorbent particles. For typical groundwater flow velocities and a

thickness of the sorptive wall of about 1-2 m, the mean residence time of contaminated groundwater will

range from less than a day to a few days. If the mean residence time is too short for equilibration, the K d

values will be much smaller than expected for equilibrium conditions, which would in turn result in an early

breakthrough of the contaminant through the sorptive wall.

Column Tests with Contaminated Groundwater From the Site

A natural bituminous shale sample and three different GACs were used in laboratory column tests in order

to investigate the transport of PAHs within the gates. Breakthrough of PAHs in the bituminous shale column

was already observed after approximately 10 displaced pore volumes, indicating that the equilibrium

sorption capacity was not reached (by far). In the GAO columns, no breakthrough was observed, and PAHs

were concentrated in the first few centimeters of the packed bed (Table 2), as determined by methanol

extraction of thin GAO layers after the column test.

The results obtained for the GAO indicate that sorption at these relatively slow flow velocities is reasonably

close to equilibrium (typical flow velocities in drinking water treatment are much higher) and the retardation

factors to be expected are higher than 3,000.
Treatment Walls NATO/CCMS Pilot Study Phase III

60

Table 2. Column tests for removal of PAHs (16 EPA PAHs) from contaminated groundwater (PAH

concentration approximately 2500 μg/l, predominantly naphthalene and acenaphthene)

Adsorbent (GAC) F100 D15/1 C40/4 (pellets) Bit. Shale

Origin Chemviron Lurgi Carbo Tech Lias & (Aalen)

Grain Size [mm] 0.5-3 1.25 4.2 2-4

BET-Surface Area m /g 770 1,102 1,281 16 2

Column Length [cm] 14 12 13 21

Porosity (intergranular) 0.52 0.68 0.61 0.48

Run Time (days) 23 23 23 112

Pore Volumes Displaced 885 731 814 1,097

Flow Velocity [m/day] 5.3 4.0 4.6 2.1

Mean Residence Time [h] 0.6 0.76 0.68 2.6

PAH Penetration Depth <1.5 cm <1.5 cm <4 cm breakthrough

Batch Experiments on Sorption of PAHs onto GAC

Batch experiments were performed in order determine the equilibrium adsorption isotherms and the sorption

kinetics of selected PAHs (acenaphthene, fluorene, phenanthrene, fluoranthene, and benz(a)anthracene) for

three different GACs and a O -modified silica gel used as reference sorbent (Table 3). 18

Table 3. Properties of GACs Used in Batch Experiments.

Property/GAC F100 C40/4 TE143 C18

Origin Bituminous Coal Bituminous Coal Coconut Silica Gel

Manufacturer Chemviron Carbo Tech Pica IST

BET Surface Area

[m /g] 2 770 1280 993 Foc = 0.19

Total Pore Volume

[cm /g] 3 0.422 0.630 0.483 -

Average Pore

Radius [Å] 10.95 9.87 9.71 -

Radius [mm] 0.8-1.0 1.0-2.0 0.4-0.5 20-35μm

K d and K , Fr determined in the equilibrium sorption experiments (Table 4) agree reasonably well with

literature data shown in Figure 2. The K d values measured in the sorption kinetic experiments were much

lower than expected from the equilibrium values (Table 4). The sorption kinetic data agreed very well with

the numerical solution of Fick’s second law for intraparticle diffusion (Figure 3). The apparent diffusion

coefficients were between 1 x 10 and 1 x l0 m s . -12 -13 2 -1

Table 4. Freundlich coefficients (K , Fr 1/n), equilibrium K d and K d values measured in the batch kinetic

experiments after 24 h, 3 days and 7 days (K *). d

F 100 C 40/4 TE 143 C18 Silica

Ace Fln Phe Ace Fln Phe Ace Fln Phe Ace Fln Phe

Log K 6.2 6.1 5.9 6.0 5.9 5.7 5.2 5.6 5.7 4.8 5.0 5.2 Fr

-1/n 0.59 0.81 0.56 0.87 0.99 0.70 0.92 0.67 0.56 0.81 0.82 0.79

Log Kd at 10% S 5.2 5.6 5.0 5.6 5.8 5.1 5.0 4.9 4.7 4.3 4.6 4.8

Log Kd* 24 h 3.4 3.7 4.1 3.2 3.2 3.3 3.3 3.3 3.4 4.6 4.9 5.1

Log Kd* 3 days 4.0 4.3 4.4 3.6 3.7 3.7 4.5 4.4 4.2 4.6 4.9 5.1
Treatment Walls NATO/CCMS Pilot Study Phase III

F 100 C 40/4 TE 143 C18 Silica

61

Log Kd* 7 day5 4.0 4.5 4.7 3.9 3.9 3.9 4.8 4.8 4.6 4.7 5.0 5.3

Hydraulic Aspects of Funnel-and-Gate Systems

There is a basic difference between the hydraulic design of a funnel-and-gate system and a continuous

permeable wall. For the continuous permeable wall, the only essential hydraulic design criterion is the

hydraulic conductivity of the wall material, which should be more permeable than the aquifer material. In

the case of a funnel-and-gate system, the hydraulic design depends on a number of factors, including the

number, position, and size of the gates, the length and angle of the funnel, and the hydraulic conductivity

of the reactive material within the gates. Starr and Cherry (1994) provided the basic functional relationships

for one-gate systems. From a practical perspective, the major design criterion is the width of the capture

zone (treatment zone) achieved for a given length of the funnel and the gate(s). The width of the capture

zone can be approximated from the total flow rate through the gate(s). In Table 5, the numerical results

obtained from the feasibility study (IMES, 1996) are summarized. It is obvious that the number of gates

(equally distributed over the length of the funnel) is of much greater influence for the width of the capture

zone than the total length of the gates.

Table 5. Dependence of the total flow rate through the gate(s) on the number

of gates and the total length of the gate(s)

Number of gates Total length of gates [m] Total flow rate [1/s]

1 12 1.04

1 36 1.55

2 12 1.58

2 24 2.01

3 18 2.09

3 24 2.22

Another aspect that needs to be considered in the design of funnel-and-gate systems is the heterogeneity of

the aquifer. The location of the capture zone can be affected by low and high permeability zones in the

subsurface. Figure 4 shows results from the 3-dimensional stochastic groundwater flow model. The three

examples (one homogeneous case and two stochastic realizations) show that the size of the capture zone

does not change even though the parameters vary considerably. However, the position of the capture zone

(little elongated box) is different for every case. The stochastic model study showed that the length of the

funnel had to be increased by a 20% safety factor in order to compensate the uncertainty related to the

heterogeneity of the aquifer.

It is also interesting to note that a tenfold increase of the conductivity of the gate only resulted in a less than

10% increase of the flow rate through the gates for a given funnel-and-gate configuration. This finding was

found to be true for gate permeabilities exceeding the aquifer permeability.

Technical Construction of the Funnel-and-Gate System at the Site

For constructing the funnel several different possibilities, e.g., slurry wall or sheet pile wall etc. are readily

available. These techniques are state-of-art and have been applied in the construction business many times.

The innovative part of the funnel-and-gate system is the construction of the gate(s) and the connection

between the funnel and the gate(s). Since it is anticipated that the sorptive activated carbon fill of the gate(s)

has to be exchanged from time to time, the gate(s) will be reinforced concrete buildings, which can be
Treatment Walls NATO/CCMS Pilot Study Phase III

62
Treatment Walls NATO/CCMS Pilot Study Phase III

63

Figure 5. Conceptual horizontal and vertical sections of the gate

construction pit

emptied. In Figure 5 the conceptual

design of the gate construction is

illustrated In a horizontal (A) and two

vertical sections (B) and (C). The pit

will be excavated after installing a

stiffened sheet pile caisson. The

concrete building will be constructed

either in place of will be prefabricated.

Simultaneous to the construction of the

gate(s), a gravel pack will be filled

between the gate building and the sheet

piles at the inflow and outflow sides of

the gate(s). After completion of the gate,

construction the sheet piles will be

pulled out, allowing the contaminated

groundwater to flow through.

With the upper two figures (“1” and

“2”) in Figure 6, two conceptually different possibilities of how contaminated

water may flow through the gates are

demonstrated: (1) horizontal flow and

(2) vertical flow. Vertical flow would be

favorable from sorption considerations.

It can be expected that the PAHconcentrations will be different at

different depths. Mixing due to vertical

flow within the gravel pack will level

out the concentration profile before the

groundwater enters the gate. Horizontal flow would be favorable from hydraulic considerations. Owing to

the aquifer geometry, the cross-sectional flow area is much larger in the case of horizontal flow, leading to

a smaller hydraulic pressure build-up upstream of the gates.

The lower two figures in Figure 6 give

some details about how the funnel (e.g.,

slurry wall) may be connected to the

gates.

Conclusions and Further Plans

Granular activated carbon appears to be

applicable for in situ removal of PAHs

occurring in the groundwater at the

former manufactured gas plant site in

Karlsruhe. For the preliminary design of

a funnel-and-gate system at the site with

two gates and about 400 m total gate 3 

volume the sorption capacity of the

activated carbon fill can be expected in

the order of 8 to 12 years.
Treatment Walls NATO/CCMS Pilot Study Phase III

64

Numerical flow modeling showed that the PAH-plume can be captured by a two-gate system with a total

gate length of 24 m. From the modeling results it also obvious that multi-gate systems are advantageous over

single or two-gate systems.

Heterogeneity structures with regard to hydraulic conductivity may have a major influence on the position

of the capture zone relative to the location of the funnel-and-gate system.

In order to work out the final design of the funnel-and-gate system, additional experimental and numerical

work will be necessary, including the performance of sorption column experiments at the site simulating

the conditions of in situ activated carbon filtration, laboratory experiments to optimize the choice of the type

of activated carbon to be used with regard to sorptive and hydraulic considerations, numerical stochastic

modeling in order to determine the length and location of the funnel and also to determine the number and

geometry of the gates.

These investigations will be carried out during the first half of 1998 and will provide the knowledge required

for the design and engineering of the funnel-and-gate system. The construction of the funnel-and-gate

system is expected to be carried out in 1999.

Figure 6. Design possibilities for the gates (1, 2) and for the

connection between funnel and gates (3,4)

References

American Water Works Association (1990): Water Quality and Treatment, 4 ed. th

IMES GmbH (1996). Machbarkeitsstudie Funnel-and-Gate am Standort Gaswerk-Ost in Karisruhe -

unpublished report.

Sontheimer, H., Frick, B.R., Fettig, J., Hörner, G., Hubele, C., Zimmer, G. (1985). Adsorptionsverfahren

zur Wasserreinigung.- DVGW-Forschungstelle am Engler Bunte Institut der Universitat Karlsruhe

(TH).

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.
Treatment Walls NATO/CCMS Pilot Study Phase III

65

Teutsch, G., Grathwohl, P., Schad, H., Werner, P. (1996). In situ Reaktionswände - ein neuer Ansatz zur

passiven Sanierung von Boden- und Grundwasserverunreinigungen. Grundwasser, 1, 12-20.

Trischler und Partner (1996). Sanierung ehemaliges Gaswerk-Ost in Karlsruhe - unpublished report.

Discussion

Noting that carbon is an excellent medium for bacterial growth, Shad was asked how to prevent bacterial

slime from clogging the system. He said that the system is anaerobic and must be maintained under

anaerobic conditions. He has performed laboratory column tests at 20-25EC and did not observe bacterial

growth. He also plans to perform sorption column tests under in situ conditions, using the same water, to

further examine the potential for bioclogging.

Wolfgang Wüst asked about the sum of the dissolved organic carbon in the system. Schad was not sure, but

said that there is highly organic material (including humic acids) present. Dissolved organic carbon will be

the main parameter examined during upcoming column tests. Schad will be assessing co-sorption of

contaminants and variations in sorption rates.
Treatment Walls NATO/CCMS Pilot Study Phase III

 Golder Associates (UK) Ltd, 54-70 Moorbridge Road, Maidenhead, Berkshire, SL6 8BN, England, +44-1628-771731, 1

e-mail: sjefferis@golder.com

 Nortel Ltd, Oakleigh Road South, New Southgate, London N11 1HB, England +44-181-945-3556, graham.norris@ 2

nt.com

66

Reactive Treatment Zones: Concepts and a Case History

Stephan A. Jefferis and Graham H. Norris 1 2

ABSTRACT

The concept of the reactive treatment zone whereby pollutants are attenuated as they move along a pathway

in the ground has enabled a re-thinking of many of the concepts of containment. In particular it allows the

control of the flux from a contaminated area by controlling the contaminant concentration in the pathway(s)

as well as or instead of using a low permeability barrier. This paper outlines the basic concepts of the

reactive treatment zone and the use of permeable and low permeability reactive systems. It then gives a case

history of the design, installation and operation of a reactive treatment zone using an in situ reaction

chamber.

1. INTRODUCTION

The use of containment systems is now widely accepted for municipal and toxic waste landfill sites but up

to now it has been more reluctantly accepted for the remediation of contaminated land. This reluctance

perhaps stems from a desire to use process orientated treatment technologies as a permanent means of

dealing with contaminated land problems. However, as the effectiveness of some in situ treatment technologies remains in doubt and the costs of treatment in general remain relatively high, interest in containment continues to develop and particularly in combining treatment with containment and the prospects for

using intrinsic processes as low cost longer term remedial options (Bardos and van Veen 1996).

1.1 Terminology

In the UK, it is common practice to refer to barrier systems that impede groundwater flow as passive

containment systems, and the term “active containment” has been coined for those that reduce pollutant flux

by treating the contaminant(s) in a flow path between a source and a receptor. However, these terms are not

universally accepted. For example, in Canada and the USA, “active containment” is often used to describe

processes where there is an energy input, such as pump-and-treat systems, and the term “permeable barrier”

is used for the UK’s “active containment.” In this paper the more general term “reactive treatment zone” will

be used for in-ground systems involving reactions that tend to reduce contaminant concentrations in a

pathway, and “passive containment” for barriers designed to impede flow.

2. REACTIVE TREATMENT ZONES

The essential feature of a reactive treatment zone is that contaminants can be controlled in a pathway

without necessarily preventing the flow of the carrier fluid (almost always water but gases are also amenable

to such treatment). 

Reactive treatment zones can be distinguished from passive (low permeability) containments by considering

the controls on the flux, which for any contaminant may be written as:
Treatment Walls NATO/CCMS Pilot Study Phase III

67

Flux = Concentration x Permeability x Hydraulic gradient x Flow area

For a traditional passive barrier, the flux is controlled by ensuring that the permeability of the system is low

and hence that the flow through the barrier is low. However, as a result the flow area will be high as the

barrier must surround the source and the concentration(s) of the contaminant species and the hydraulic

gradient (unless modified by pumping) also may be high.

An alternative way to control the flux is to reduce the concentration of the contaminant(s). If a sufficient

reduction in concentration can be achieved then permeability ceases to be the controlling parameter and it

may be high or low as appropriate to the application. With a high permeability system the hydraulic gradient

will be low as the carrier fluid is allowed to escape and the area for flow may be less than that of a containment as it may be deliberately focused with an engineered system or naturally higher permeability regions

may be exploited—if they can be sufficiently well defined. This may simplify monitoring, and where

necessary, in situ control. If space is available the input concentration may be varied by selecting the position

of the reactive zone in relation to the contaminant source or in-ground intrinsic remediation may be

exploited upstream or downstream of the reactive treatment zone. It should be noted that reactive treatment

zones often will be used as pathway control mechanisms to prevent contaminants from a source reaching

a receptor rather than as source clean-up technologies (at least in the short term).

2.1 Reactive Zone Configurations

In principle a reactive treatment zone or zones should completely enclose identified pathways from the

pollutant sources to the receptors. In practice this may require that the zone(s) cover a substantial area.

Furthermore the flow through any element of a zone may be uncertain as a result of factors including

seasonal variations in groundwater level and flow. It follows that large volumes of reactive material may

be necessary to ensure sufficient residence time throughout the zone under all operating conditions.

Optimization of the deployment of the reactant is therefore an important consideration. Possible

configurations include:

 • The use of injection wells to introduce chemicals into the natural groundwater flow.

 • The use of passive wells containing, for example, replaceable canisters of treatment materials that

dissolve into the natural groundwater flow.

 • Treatment zones in the region of the source that reduce pollutant concentrations to levels at which natural

intrinsic remediation can operate in zones down gradient of the source.

 • Systems that focus the flow either in plan or elevation or both.

It should also be possible to monitor and, if appropriate, control the processes within the treatment zone. The

requirements for monitoring and control can present problems for in situ treatments or intrinsic remediation

in an undefined area, given the heterogeneity typically associated with ground conditions and the large

volumes of soil that may require treatment. Martin and Bardos (1996) describe examples of in situ treatment

zones where treatment remains in the ground but under conditions of better definition. The gate of a Funnel

and Gate system is an example of such a system where monitoring can be particularly straightforward and TM

control systems can be incorporated in the works if appropriate.

2.2 The Permeability of Reactive Treatment Zones

Reactions in a treatment zone may influence its permeability to the carrier fluid just as they may influence

the permeability of a passive containment. It follows that chemical, biochemical and physical phenomena

within contaminant control systems must be recognized as interrelated. Contaminant control systems should

be considered as a continuum bounded by the two extreme situations:
Treatment Walls NATO/CCMS Pilot Study Phase III

68

• Permeable reactive treatment zones. The best known example is the use of zero valent iron to

dechlorinate solvents in groundwater (Gillham and O’Hannesin, 1992).

• Low permeability active containments, achieved, for example, by exploiting the chemistry of, or adding

species to, low permeability barriers such as a soil-bentonite (Evans, 1991) or cement-bentonite (Jefferis,

1996).

A reactive treatment zone must include some process that removes, destroys precipitates or otherwise

attenuates the contaminants migrating from the source(s) towards the receptor(s). Processes that may be

employed alone or in combination include: chemical reaction, physical separation, biological degradation,

and sorption. 

Many of these processes will rely on an interaction between the flowing groundwater and a solid phase in

the treatment zone and although the permeability of the zone is not the controlling parameter it may be

necessary to consider it. For example, a moderate to high permeability may be necessary if problems such

as ponding within a containment are to be avoided. Designing and maintaining such a permeability may pose

some problems as the reaction rate in the zone may be dependent on the surface area of a solid phase within

it. As a result it may be necessary to use finely divided (and hence low permeability) materials to limit the

required residence time of the contaminants in the zone. Also precipitates or biological slimes may form

within the zone. Thus careful consideration must be given to the initial permeability of the treatment zone

and to processes that may lead to any change over time (an increase as well as a decrease in permeability can

be damaging if the water level or residence time in the reactive zone is sensitive to its permeability). It is

perhaps ironic that a low permeability barrier may be compromised by an increase in its permeability

whereas a reactive treatment zone may be compromised by a decrease. The problems are reciprocal and if

one type of containment could be perfectly achieved there would be no need for the other.

2.3 Low Permeability Reactive Zones

Reactive treatment zones may be designed to be of low permeability, and the physical effect of the barrier

may be enhanced by processes such as chemical reaction in the region of the barrier. For example, heavy

metals may be precipitated by the high pH of cement-bentonite walls though credit is seldom given for this

in the design of such containments (except for the containment of radionuclides in purpose-designed

repositories). 

A significant amount of work has been reported on the effect of contained chemicals on the physical

properties of barrier materials such as clay liners and cement or soil-bentonite walls. The prime focus of

much of this work has been the assessment of physical damage to the cut-off material, for example, its

permeability that may be caused by contained chemicals. Much less work has been reported on the beneficial

effects of chemicals, though an examination of the literature shows that some work focused on the damage

has actually demonstrated beneficial effects.

Interactions in low permeability systems are potentially much more complex than in high permeability

systems as it is necessary explicitly to consider the effects of chemical, biochemical and mineralogical

changes on the permeability and not just design to minimize their impacts. At the laboratory scale it has been

demonstrated that the chemical history of a containment material and its permeability can be strongly

interlinked. For example, leaching a cement-bentonite material by sustained permeation with water, so

reducing its pH, will usefully reduce its permeability and also inhibit expansive cracking and damage if it

is subsequently exposed to sulfates. High pH, if required, can be re-established by permeation with lime and

this can further reduce the permeability but it will re-establish sensitivity to sulfates (Jefferis 1996). 
Treatment Walls NATO/CCMS Pilot Study Phase III

69

That such interaction can occur is not surprising but it does raise questions about the interpretation of

laboratory investigations of chemical damage to barriers using samples that have not been exposed to the

chemical history that occurs in situ (including, and very importantly, periods of leaching by water). Also it

has been demonstrated that some chemical damage can be reversed by re-introduction of appropriate

chemical species. It follows that there is considerable potential to modify the behavior of barriers once in

the ground. This could be to repair them, extend their life, protect them from damaging species or modify

or re-charge their chemical activity. It is therefore quite simplistic to consider cement-bentonites and indeed

most passive barrier systems as having a unique permeability. They can be physically, chemically or

biochemically active and if we can learn to exploit this activity new barriers concepts will be forthcoming.

3. CASE HISTORY OF THE APPLICATION OF A PERMEABLE REACTIVE ZONE

In addition to the problems of design of reactive treatment zones with appropriate chemical reactions there

can be significant practical problems (and also opportunities) in fitting such systems to the circumstances

of a site and this is illustrated by the following case history.

At an industrial site in Belfast, Northern Ireland, currently being used for the manufacture of electronic

components, historic spillages of chlorinated solvents had led to an intense though localized contaminant

source and plume. A site investigation identified the following contamination in the groundwater:

 • Trichloroethene (TCE) at concentrations up to 390 mg/l

 • Trichloroethane (TCA) at concentrations between 100 and 600 μg/l

 • Tetrachloroethene (PCE) at concentrations of the order of 100 μg/l

A particular problem with chlorinated solvents is that they tend to be of low solubility (typically less than

about 1,000 mg/l) and thus can remain as a free phase product in the ground for substantial times but the

regulatory control levels are typically of the order of tens of mg/l—i.e., of the order of one hundred thousand

times less than their solubility. Also they are generally recalcitrant and not easily biodegraded. Thus even

quite a modest chlorinated solvent spill may persist in the groundwater and continue to pollute it for a

substantial time.

The soil profile at the site was complex, with glacial till interbedded with lenses of silts, sands and gravel.

in situ pumping tests showed that the till acted as an effective barrier to the vertical and lateral migration of

the contaminants but the silt, sand and gravel lenses were sufficiently permeable to allow off-site migration

of dissolved solvents from the source that included free phase solvents present as a dense non-aqueous phase

liquid (DNAPL). There were also smaller discrete lenses of clay or clayey silt that complicated the flow

regime.

In situ permeability tests and pumping tests conducted to evaluate potential remedial strategies showed that

conventional pump and treat remediation would require numerous extraction wells to address the thin

saturated zone and this, together with the presence of a free product DNAPL source, meant that such a

system would have to be maintained for many years. Similarly there would be major problems with a soil

vapor extraction system.

3.1 Reductive Dechlorination

One of the best known and researched reactive treatments is the use of iron filings to reductively

dechlorinate chlorinated solvents (Gillham and O’Hannesin, 1994). The process essentially leads to the

development of an extremely reducing environment, allowing reactions occurring on the surface of the iron

to strip halogens from the dissolved organic compounds as they flow through the zone. The reaction is

abiotic and chlorinated solvents are converted to chloride ions and various hydrocarbons.
Treatment Walls NATO/CCMS Pilot Study Phase III

70

The process does not degrade the resulting hydrocarbons but as non-chlorinated species they have much

higher permissible regulatory control concentrations. Also they are much more readily bio-degraded by

intrinsic remediation processes occurring in the ground. The iron filings reactive containment is thus a

conversion process that reduces toxicity. This demonstrates a general principle that a reactive treatment zone

need not degrade organic contaminants to basic species such as carbon dioxide and water, but need only

convert them to species that are acceptable toxicologically and environmentally in the local setting.

The iron filings technology seemed to have potential for the Belfast site and samples of the site groundwater

obtained from sampling wells were shipped to EnviroMetal Technologies, Inc., Guelph, Ontario, for

treatability studies. The results of these studies are reported in Thomas et al. (1995). Briefly, a series of

laboratory column studies were undertaken to determine the degradation kinetics of the species of interest

and to identify whether there were any constraints imposed by the natural geochemistry of the site

groundwater. The results showed that TCE, PCE and TCA were completely degraded from initial

concentrations in the test groundwater samples of 300,000, 170 and 200 μg/l respectively. The half life of

TCE in the site water was estimated to be 1.2 hours in the presence of iron filings.

In contrast small concentrations of cis-dichloroethene (DCE) and vinyl chloride (VC) built up in the water

as a result of the dechlorination of the TCE, etc., with VC reaching a peak at the end of the column of 700

μg/l in the initial studies. The reason for this is that although the dechlorination does not appear to be a

stepwise process (i.e., tetrachloroethene to tri- and then dichloroethene, vinyl chloride, and ultimately to

chlorine-free hydrocarbons), and it seems that the majority of the solvent is dechlorinated in a single step.

A small amount may be released from the surface of the iron prior to full degradation. This leads to a

temporary increase in species, such as DCE and VC, as these species are more slowly dechlorinated by iron

filings. Considerable care was necessary in the design to ensure that the DCE and VC would be satisfactorily

degraded. Otherwise, the net effect of the treatment zone could be to increase the concentrations of these

species, which would be most unsatisfactory, particularly as vinyl chloride has markedly lower regulatory

control limits than many other chlorinated solvents.

Further column studies were then undertaken at lower flow rates, the increased residence time resulting in

much earlier degradation of the parent compounds thus enabling estimation of the degradation rate of vinyl

chloride. From these laboratory trials, it was concluded that a residence time in contact with the iron filings

of about 12 hours was required for the site water to achieve removal of all chlorinated species to below

regulatory limits. This result, together with the results of a groundwater modeling exercise, enabled the

calculation of the required reactive zone path length from the estimated groundwater flow rate and the

required residence time of the contaminated groundwater in contact with the iron filings.

3.2 Design of the Reactive System

A preliminary review of the site situation concluded that the optimum strategy for deployment of the iron

filings was in the gate of a Funnel and GateTM system. However, the site geology and site circumstances

placed a number of constraints on the design:

 • The contaminant source extended to within a few meters of the site boundary, outside which there was

a public road. The reactive treatment zone therefore needed to be very compact.

 • The solvent source was underlain by a thin layer of clay, which had inhibited its migration to greater

depths. If this layer were penetrated by the gate the free product solvents would sink and pollute a lower

aquifer stratum.

 • The groundwater perched on the thin clay layer was shallow and showed seasonal variations in depth.

It would be difficult to maintain any significant depth of horizontal flow in a gate seated on this layer.
Treatment Walls NATO/CCMS Pilot Study Phase III

71

Figure 1. Reactor configuration, elevation,

and plan view

 • A perched water table also existed in the fill covering the surface of the site. In wet seasons this water,

if allowed to enter the gate, could dominate flow through it and unacceptably reduce the residence time.

It therefore had to be excluded.

 • Because of the proximity of buildings and the risk of damage by vibration and also the cost, sheet piles

could not be used to form either the funnel or the gate of the system.

 • If a slurry trench procedure were used to form the system then it was imperative that the iron filings were

not inundated and thus blocked by slurry from the wall construction. It would therefore be necessary to

contain the iron or to construct the slurry wall before placing the iron filings. Prior formation of the cutoff wall could allow it to slump during excavation for the iron gate and thus a poor seal between the wall

and the iron filings.

 • The residence time of groundwater within the reactive treatment zone is critical. A zone that is too thin

may lead to the escape of lower chlorinated species, such as vinyl chloride, which can pose a greater

toxicological or environmental threat. Extensive groundwater modeling studies were undertaken to

ensure that the design residence time could be achieved. 

 • The clean-up was being undertaken voluntarily by the owner of the site.

After consideration and rejection of many reactive treatment zone designs, the in situ reactor configuration

shown in Figure 1 was developed as best fitting the constraints imposed by the site. In place of previously

used horizontal flow reactive treatment zones the flow was arranged to be vertical in a 12 m tall by 1.2 m

diameter steel reactor shell which was filled with iron filings as shown in the figure. This design enabled

the reactor to be placed between the contamination

and the site boundary. This could not have been

achieved with a horizontal flow regime as the design

calculations had shown that the flow path length

needed to be at least 5 m plus entry and exist zones

to collect and disperse the flow. The vertical flow

direction within the reactor also ensures that the full

depth of the iron filings will be saturated whatever

the seasonal variation in groundwater level.

The reactor was placed in an enlargement in a

cement-bentonite cut-off wall (Figure 1), which was

used to funnel the flow to the reactor. This wall was

toed into an aquiclude layer at 10 m and the enlargement was taken to a depth of slightly over 12 m to

accommodate the reactor shell. The cut-off and

enlargement penetrated through the clay layer on

which the chlorinated solvents were retained.

However, as the cut-off material was designed to

have a permeability of <10 m/s minimal downward -9 

migration of solvents will occur.

Because of the relatively low permeability and

heterogeneity of the adjacent soil, it was decided that

the flow to the reactor should be collected via an

upstream high permeability collector and that

downstream of the reactor there should be a similar
12-Aug-97

21-Jan-97

24-Oct-96

17-Jul-96

18-Apr-96

0

10,000

20,000

30,000

40,000

50,000

60,000

70,000

80,000

Treatment Walls NATO/CCMS Pilot Study Phase III

72

Figure 2. TCE concentrations in the reactor zone

distributor. The collector and distributor were formed from gravel filled piles taken down to the top surface

of the thin clay layer. A gravel backfilled trench, excavated under a degradable polymer slurry, could have

been used to form the collector and distributor but there were concerns about the effect, on the iron filings,

of any polymer remaining in the trench and so augured pile holes were used.

Finally, the internal geometry of the reactor was arranged so that the pipe connections to the gravel filled

collector and distributor piles could be made from within the clean environment of the reactor shell without

the need for any hand excavation or for anyone to enter the excavations all of which were undertaken either

with a backhoe under cement-bentonite slurry or with a piling auger. During the works airborne solvent

concentrations were monitored and found to be undetectable.

3.3 Performance of the Reactor

The performance of the reactor has been and continues to be very satisfactory though there are two marginal

complicating factors:

 • The flow rate through the reactor was very low during the first months of operation and is likely to

remain quite modest (this is as expected from the hydrological investigation). It follows that the internal

pore water chemistry is rather sensitive to any disturbance of the flow regime

 • Sampling from within the relatively modest volume of the reactor has proved to be a problem if

traditional procedures are used. Only small volumes of water can be withdrawn if the local chemistry

within the reactor is not to be disturbed. Particular techniques are being developed to obtain small but

representative samples from the reactor whilst avoiding the need to purge the sampling tubes.

The results of five campaigns of sampling within the reactor and in the upstream collector and downstream

distributor are shown in Figures 2, 3, and 4. The results relate to a period of 19 months from installation of

the reactor. 

Figure 2 shows the data for TCE as a

three dimensional plot of concentration

as a function of position in the reactor

system and time. There are five sampling

points within the reactor shell. Sampling

point R5 is at the entry point to the iron

filings bed from the collector and R1 at

the discharge point to the distributor.

The other three sampling points, R2

through R4, are distributed at 1.5 m

intervals through the 6 m deep bed. 

It can be seen that the of TCE within the

reactor drops very rapidly and is at very

low levels by the time it has reached the

first sampling point (R4) at 1.5m into the reactor bed. In August 1997 the average TCE concentration within

the reactor bed was less than 4 micrograms/liter. Figure 2 also shows that there is still a significant, though

reducing, TCE concentration in the downstream distributor and the reactor discharge point (R1). This is as

a result of the TCE originally present in the downstream area some of which may have diffused back into

the reactor exit zone or have been drawn back into the reactor by the sampling process. However, it seems

most likely that the actual reason for its presence is the advection of TCE contaminated groundwater through
0

50

100

150

200

250

300

Collector R5 R4 R3 R2 R1 Distributor

Location

TCE

18-Apr-96

17-Jul-96

24-Oct-96

21-Jan-97

12-Aug-97

Total

18-Apr-96

17-Jul-96

24-Oct-96

21-Jan-97

12-Aug-97

Treatment Walls NATO/CCMS Pilot Study Phase III

73

Figure 3. Chlorine balance in the reactor zone

the distributor zone as a result of groundwater flow in the downstream area. It should not be forgotten that

there may be a downstream flow vector parallel to as well through a gate. 

Figure 3 shows the chlorine balance in the reactor zone. For this it is necessary to consider both the chlorine

in the inorganic chloride ion and in the TCE. The TCE chlorine drops rapidly with distance into the reactor

and only increases again in the distributor area. The total chlorine shown in the figure is the sum of the TCE

and chloride chlorine and it can be seen that it is relatively consistent throughout the reactor (the high total

chlorine on 21 August 1997 was as a result of a high chloride ion concentration, the reason for it is not

known but could be as a result of some slight seepage of de-icing salts used in winter time.

On each sampling occasion TCA was at undetectable levels. Vinyl chloride was found at a single sampling

point within the reactor at 4 months (at a level of 0.4 mg/l) but not in the collector or distributor and has not

been detected on any subsequent occasion.

With regard to the remaining downstream contamination it is interesting to note, that at times of low

seasonal flow, there is potential for the recycling groundwater from wells downstream of the reactor to the

upstream collector (by pumping) to treat the downstream area should treatment of this contamination be

required. Also the reactor will work with the flow in either direction. Thus, if, for some reason, the flow

should be reversed naturally then the site would be protected from the inflow of any contaminants spilled

or remaining outside the reactor system.

4. CONCLUSIONS

The reactive treatment zone concept offers not only a new tool for the control of contamination but also

forces a re-thinking of low permeability barrier systems. Prior to the development of the concept,

contaminant/barrier interaction had generally been considered only as a possibly damaging process that

could increase the permeability of a barrier. Now barrier interactions should be seen differently. They may

be deliberately engineered to be beneficial.
Treatment Walls NATO/CCMS Pilot Study Phase III

74

Reactive treatment zones may be most economically employed as pathway control procedures but will

contribute to the slow clean-up of a source. The pathway control they offer is their essential feature and it

may be argued that they should not be designed to achieve more than slow source clean-up as this minimizes

the rate at which any treatment chemicals, energy or other process inputs are required i.e. it minimizes the

process demands and operating costs whilst maximizing the opportunity for natural, intrinsic, remediation.

At the present time a significant limitation to the use of reactive treatment zones is the rather narrow range

of contaminants that can be treated and the complications introduced by mixed contaminants etc. However,

work is in progress and the technology has strong parallels with intrinsic remediation (e.g., intrinsic

bioremediation), and it is likely that the reactive treatment zone concept will find much application in

stimulating intrinsic degradation or attenuation.

The Funnel and Gate concept is a powerful adjunct to the technology as it can allow the treatment zone TM

to be at some distance from the source, if this is required by surface or buried structures, geological

conditions or for process optimization (e.g., exploiting natural dilution or attenuation in the soil), etc. Sites

need not be capped to prevent rainwater ingress, indeed it can be an advantage in driving groundwater flow

to a treatment zone.

The concept of an in situ reactor adds further flexibility to the design of reactive treatment zones, allowing

more precise control of the reactive zone and also recharging or recovery and replacement of the active

material should this be required. Furthermore several reactors may be linked in series to treat mixed

contaminants.

Although most reactive treatment systems that have been advanced to date have been for the control of

groundwater pollution there is no reason why other systems such as reactive caps should not be developed

to control, for example, the escape of semi-volatile or volatile organics from petroleum impacted sites or gas

or odor from landfills. Peat or humic based materials could be particularly cost effective in this respect (and

of course for the removal of organics from impacted water). Bio-oxidation zones to control methane

migration through unsaturated soils offer some exciting prospects but there could be difficulties with

excessive local heating. 

Low-permeability reactive treatment zones are a more specialist application and may require site capping

or groundwater pumping (and treatment) to avoid ponding. Reactive treatment zone concepts will be useful

where long term secure containment is necessary as chemical interactions can provide additional security

and lifetime to a physical containment. Furthermore the concepts may be useful in developing in situ repair

strategies for barriers that have suffered damage and in analyzing the effects of reaction on the permeability

of these systems (Jefferis, 1996).

5. ACKNOWLEDGMENTS

Golder Associates gratefully acknowledges the proactive technical and financial support given by Nortel Ltd

who funded the reactive treatment project referred to in the case study.

The support of the UK Department of the Environment, Contaminated Land and Liabilities Branch, is also

gratefully acknowledged. They have funded Golder Associates to undertake a research project EPG 1/6/16

on active containment concepts. This paper includes review work carried out for the Department of the

Environment. However, the views expressed in this paper are those of the authors only and do not

necessarily reflect the views of the Department of the Environment. Reference to any individual or

organization or mention of any proprietary name or product in this paper does not confer any endorsement

by the authors nor the Department of the Environment.
Treatment Walls NATO/CCMS Pilot Study Phase III

75

The award of an ERDF Northern Ireland Single Programme (1994 - 1999) Environmental Services and

Protection Sub-Programme grant of 50% of qualifying costs associated with the Nortel reactive barrier

project is also gratefully acknowledged.

6. REFERENCES

Bardos, R.P. and J. van Veen. (1996) Longer term or extensive treatment technologies. Land Contamination

& Reclamation, 4(1), pp. 19-36.

Evans, J.C., Y. Sambasivam and Z.R. Zarlinski. (1991) Attenuating materials in composite liners, Waste

Containment Systems: Construction, Regulation and Performance, ASCE Geotechnical Special

Publication, No 26. 

Gillham, R.W. and S.F. O’Hannesin. (1994) Enhanced degradation of halogenated aliphatics by zero-valent

iron, Groundwater, 32(6), pp. 958-987.

Gillham, R.W. and S.F. O’Hannesin. (1992) Metal catalysed abiotic degradation of halogenated organic

compounds, IAH Conference “Modern trends in Hydrogeology”. Hamilton Ontario, Canada, pp. 94-

103.

Jefferis, S.A. (1996) Contaminant - barrier interaction: friend or foe? Mineralogical Society Conference,

Chemical Containment of Wastes in the Geosphere.

Martin, I. and R.P. Bardos. (1996) A review of full scale treatment technologies for the remediation of

contaminated soil. (Final report for the Royal Commission on Environmental Pollution, October 1995).

EPP Publications, 52 Kings Road, Richmond Surrey.

Starr, R.C. and J.A. Cherry. (1994) In situ remediation of contaminated groundwater: The funnel-and-gate

system, Groundwater, 32, pp. 465-476.

 

Thomas, A.O., D.M. Drury, G. Norris, S.F. O’Hannesin, and J.L. Vogan. (1995) The in situ treatment of

trichloroethene contaminated groundwater using a reactive barrier - results of laboratory feasibility

studies and preliminary design considerations. Contaminated Soil ‘95, W.J. van den Brink, R. Bosman

& F. Arendt (eds).

Discussion

There was some discussion concerning the placement of the reactor vessel relative to the collector and

distributor, and whether there would be any cost savings in installing the reactor at the same time as the cutoff wall, collector, and distributor. Jefferis replied that there are circumstances where there would be no cost

savings for simultaneous construction. He said that sometimes it would be cheaper to build a pipe through

the cut-off wall rather than place the reactor within it.

The reactor was designed to have a retention time of approximately 1 day. It is difficult to determine the

exact flow through of the reactor because it was operated in the summer when the through-flow was too low

to measure. The highest flow occurs in the winter, and the reactor was designed for a maximum flow through

of 5 m /day. Other chemical parameters measured at the site were pH, PCE, DCE, vinyl chloride, and some 3

inorganics.
Treatment Walls NATO/CCMS Pilot Study Phase III

76

Harald Burmeier questioned whether the groundwater was pumped through the reactor. Jefferis indicated

that it is a gravity-flow reactor, but if necessary, groundwater could be pumped through to treat water, or to

pump down in the winter during peak flow to augment the low flow during the summer. 

Wüst commented that the system seemed to work well, but wondered whether the hydraulic design was

adequate to clean up other contaminants besides TCE. Jefferis noted that no other contaminants were

present, other than cis-DCE, which was found only in the distributor (not in the flow through). He added

that the higher concentrations shown in Figure 2 were found in the distributor located downstream of the

reactor. He said that the reactor has unused capacity, and he would like to pump back the flow through; he

acknowledged his disappointment in how slowly the concentration was dropping. Geologically, it is a very

heterogeneous, complex site. There may be some cross-flow bringing in contamination directly from the

plume. 
Treatment Walls NATO/CCMS Pilot Study Phase III

 Colorado School of Mines, Environmental Science and Engineering Division, Golden, CO. USA 80401-1887. Ph. 303- 1

273-3490. Fax. 303-273-3490. Email. rsiegris@mines.edu. 

 Oak Ridge National Laboratory, Environmental Sciences Division and Life Sciences Division, Grand Junction, CO. 2

 FRx, Inc., Cincinnati, OH and Clemson University, Clemson, SC. 3

 Lockheed Martin Energy Systems, Inc., Piketon, OH. 4

77

Horizontal Treatment Barriers of Fracture-Emplaced Iron and Permanganate Particles

Robert L. Siegrist , Kathryn S. Lowe , Lawrence W. Murdoch , 1 2 3

Traci L. Case , Douglas A. Pickering , and Thomas C. Houk 1 2 4

1.0 Abstract

In many contaminated sites, vertical leaching and/or volatilization of organic compounds can lead to high

exposures and unacceptable human health or environmental risk. One remediation alternative for these sites

is to emplace horizontal treatment barriers to reduce the source concentration and mass as well as intercept

and treat any mobile organics in the subsurface. Development and demonstration activities have been

conducted to evaluate the feasibility of creating horizontal treatment barriers by employing hydraulic

fracturing for emplacement of chemically reactive solids. In this work, iron metal particles and a new

permanganate solid media were both tested in horizontal barriers to achieve interception and degradation

of chlorocarbons like trichloroethene (TCE). Laboratory experiments were completed to develop suitable

delivery fluids and to determine reaction rates and efficiencies. The iron metal particles could be suspended

and delivered using a typical guar gum gel while the permanganate solids required a new mineral-based

carrier fluid. Degradation of TCE at ~10 to 80% aqueous solubility with iron metal particles exhibited a halflife in the range of 1 to 2 hr while that of the permanganate solid was much faster with a half-life of only

several min or less. A full-scale field test was subsequently conducted at an old land treatment site in the

midwestern USA. Two test cells were installed with horizontal barriers comprised of reactive fractures

emplaced at ~1.2, 2.4, and 3.6-m depth in silty clay soils using hydraulic fracturing methods. After 3, 10 and

15 mon of emplacement, continuous cores were collected across the emplaced fracture zones and

morphology was observed, geochemical properties were measured, and degradation tests were completed

with TCE contaminated groundwater (TCE concentrations at ~5-50% solubility). These analyses revealed

strong short-range trends in redox potential and pH consistent with the anticipated behavior of the two

different reactive media emplaced. Highly oxidizing zones were present in and around the permanganate

barriers. TCE degradation efficiencies during 2 hr of contact of >99% in 10-cm thick soil zones after ~3 mon

of emplacement with even broader effects at ~10 and 15 mon. The treatment barrier comprised of 200 μm

iron-metal particles exhibited highly reducing conditions within the iron-metal fracture but no marked

effects in the soil zone surrounding it. Degradation of ~35% was achieved after 24-48 hr of contact but only

in the iron-metal itself. The iron exhibited some corrosion, but this did not appear to affect its reactivity. The

results of the work to date are very positive regarding the performance for both types of horizontal barriers

and the estimated costs for a typical application are comparably low at only $30/m . Further work is ongoing 3

to enable design and deployment of this technology.

2.0 Introduction

Petrochemicals (e.g., benzene) and chlorocarbons (e.g., TCE) are common and problematic contaminants

of concern (COCs) at federal facilities and industrial sites across the U.S. and abroad (USEPA, 1992; API,

1995; DOE, 1996). These contaminants can be released to the environment through leaks in storage tanks

and transfer lines, spills during transportation, and the land treatment of wastes. They are often present in

source areas and in soil and groundwater plumes as dissolved or sorbed phase constituents as well as

nonaqueous phase liquids (NAPLs). When these COCs are present in silt and clay media either as massive
Treatment Walls NATO/CCMS Pilot Study Phase III

78

surface deposits or as interbedded lenses in otherwise permeable formations, there are major challenges with

assessment of their behavior and implementation of effective in situ remediation technologies. Despite a low

bulk permeability of silts and clays (e.g., K < 10 cm/s), organic compounds can contaminate fine-grained sat

-7

deposits by preferentially moving into and through naturally occurring pore and fracture networks and

partitioning between the nonaqueous, gas, aqueous, and sorbed phases. In the unsaturated zone, organics can

continually volatilize into the soil air or leach into percolating water while in the saturated zone they can

slowly dissolve into advecting groundwater. Exposures and unacceptable risk can result from ingestion of

contaminated drinking water or surface soil media, or inhalation of vapors in ambient air or emitted during

showering events. In recognition of the severity of the problem and the need for effective in situ treatment

methods, organic compounds in fine-grained deposits was recently ranked as one of the top environmental

restoration needs across the DOE Complex (DOE, 1996). Similarly, within the petroleum industry, nearly

40% of the underground storage tanks in the world are located on clay soils and leaking underground tanks

and remediation of organic contamination in these settings remains a major challenge (API, 1995).

In situ remediation by conventional methods such as mass recovery by soil vapor extraction (SVE) or in situ

destruction by biodegradation or chemical oxidation are often ineffective for NAPL compounds at sites with

silt and clay media. Poor accessibility to the contaminants and difficulty in delivery of treatment reagents

throughout these deposits have made rapid and extensive in situ remediation difficult. Some aggressive

techniques have been developed involving subsurface disruption by soil mixing or alternative driving forces

based on electrokinetics (e.g., Siegrist et al., 1995). In seeking less intensive methods that could be used over

larger areas, soil fracturing techniques have been developed whereby hydraulic or pneumatic fluids can be

injected to create permeable horizontal layers. This can increase the overall permeability of the subsurface

and improve contaminant recovery as existing channels or pathways can be expanded or new channels or

pathways can be created (U.S. EPA, 1993; Murdoch et al., 1997). Hydraulic methods normally employ an

agent or proppant (e.g., sand) to fill and support the fracture opening that was created and thereby prevent

fracture closure during natural healing processes in unconsolidated deposits. If the fractures can be spaced

throughout the deposit, they potentially could be used to (1) enhance the recovery of organic compounds,

(2) deliver and distribute treatment fluids into the deposit and accomplish destruction in place, or (3) place

reactive solids as an integral part of the emplaced fracture that could serve as a horizontal treatment barrier.

Two reactive media that appeared appropriate for use in horizontal treatment barriers included: (1) iron

metal particles and (2) permanganate crystals. The use of zero valent iron metal as a reactive barrier media

has been developed and deployed in vertical orientations for several years (e.g., O’Hannesin and Gillham,

1998). The reductive dechlorination reaction for TCE is shown in equation 1. The reaction involves single

and multiple electron transfers following pseudo-first order kinetics (at least at lower TCE concentrations)

with half-lives on the order of 30-60 min as normalized to solution to solid surface area. The pH of the

reacting system tends to rise, but appears to stabilize in the pH 9-10 range due to iron hydroxide

precipitation.

2Fe + C HCl + 3H O 6 C H Cl + H + Cl + 3OH + 2Fe [1] o - - ++

2 3 2 2 2 2 2

The use of potassium permanganate for oxidative degradation of organic pollutants in contaminated land

has evolved over the past few years (e.g., Gates et al., 1995). The oxidative destruction of TCE is given in

equation 2. The reaction can include destruction by direct electron transfer or free radical advanced

oxidation. The reaction follows pseudo-first order kinetics and is very rapid with half-lives on the order of

sec to min. The pH of the reacting system can decline to very low values depending on the buffering capacity

of the system.

2KMnO + C HCl 6 2CO + 2MnO + 2KCl + HCl [2] 4 2 3 2 2

 
Treatment Walls NATO/CCMS Pilot Study Phase III

79

Figure 1. Illustration of a horizontal treatment barrier concept with

hydraulic fracturing for reactive media emplacement (Note: the

barriers shown contain reactive media that has the ability to

dissolve and effect soil-solution biogeochemistry changes).

3.0 Horizontal Treatment Barriers

3.1 Development and Demonstration Activities. 

While there has been extensive research and development followed by increasing application of vertical

treatment walls, research and development of horizontal treatment barriers has only recently been

accomplished (e.g., Murdoch et al. 1997, Siegrist et al. 1997). During the past three years, these authors have

been studying horizontal barriers comprised of either iron metal or permanganate solids delivered via

hydraulic fracturing methods. This work has included laboratory and field testing to evaluate their feasibility

and performance for in situ degradation of TCE in fine-grained media deposits (Figure 1). The results and

implications of this work are highlighted

in this paper while details regarding the

methods and results can be found elsewhere (e.g., Murdoch et al., 1997;

Siegrist et al., 1997; Case and Siegrist

1997). 

. 

The laboratory and field research

conducted by these authors has demonstrated that conventional hydraulic

fracturing equipment and methods could

be employed to emplace reactive solids

in silty clay soils at depths of 1.2-5 m

with diameters of 6-9 m. Fractures

containing 200 μm iron particles as the

proppant were successfully emplaced at

depths of ~1.8, 2.4, and 3.6 m. Based on

morphology and geochemical

measurements along profiles transecting

the fractures, the iron proppant remained

reactive even after 10-15 months of

emplacement, but there was little effect on the enveloping soil matrix (Figure 2). Based on batch degradation

tests with samples of the iron from the fracture or soil media surrounding it and initial TCE groundwater

concentrations ranging from ~50-500 mg/l (equivalent to ~400-4000 mg/kg), the degradation achieved was

equal to only ~35% during reaction periods of 24- or 48-hr. While the iron metal in the fracture did show

signs of surface corrosion, its degradation efficiency was comparable to that of unused iron even after nearly

a year of emplacement in the moist silty clay subsurface.

Fractures containing a new permanganate solid mixture as the fracturing fluid and proppant were

successfully emplaced at depths of ~2.1, 2.7, and 3.4 m. Morphology and geochemical measurements along

representative profiles indicated that this process created highly reactive fractures enveloped by zones of

reactive soil 30-60 cm thick (Figure 3). Dissolved TCE with initial concentrations of from ~50-500 mg/l

(equivalent to ~400-4,000 mg/kg) was completely degraded (100% efficiency) in batch tests using either

material from the fractures or the reactive soil from above and below the fractures. Complete degradation

was observed in as little as 2 hr. The degradation potential as a function of time and space, was controlled

by the mass of permanganate ion that was present in the fracture or the fraction of it that had migrated into

the soil matrix and had not been previously consumed.
Treatment Walls NATO/CCMS Pilot Study Phase III

80

Figure 2. Features of iron metal horizontal barriers 10 months after

initial emplacement in a silty clay soil (after Siegrist et al., 1997).

(a) Media redox potential and pH above and below the horizontal

treatment fracture. (b) TCE degradation potential above and below

the horizontal treatment fracture as measured in batch tests using

3-5 g of media in 40 ml of groundwater.

3.2 Discussion and Implications. 

The viability of reactive fractures

emplaced by hydraulic fracturing

methods as horizontal barrier systems

requires consideration of the horizontal

continuity, degradation capacity, and

longevity. The results obtained by these

authors enable a very interesting contrast

between reactive fractures created with

surface reactive particles (i.e., iron metal

particles) versus those created with a

reactive grout that dissolves and yields a

wider reactive zone (i.e., permanganate

grout). Both types of reactive media

were successfully handled and emplaced

by conventional hydraulic fracturing

equipment and methods. The geometry

of the fractures was similar to that of

conventional sand-filled fractures

emplaced at the same site. Thus, there

was no unusual behavior associated with

the different fracturing proppants (i.e.,

iron particles in guar gum versus

permanganate particles in mineral carrier

versus sand in guar gum). 

Since the iron metal fractures are

discrete reactive sheets, it is not likely

that these fractures would be self-healing

within or between fractures or effective

beyond the boundary of initial

emplacement. This challenges the fracture emplacement to be continuous and uniform horizontally with no

breaches through it. Also, degradation of the TCE or related compounds within the deposit must rely on their

being mobilized to the fracture where they can contact the metal surface. Since the TCE reaction rates with

iron metal are comparatively slow, any mobile contaminants must remain in contact with the iron metal for

a reasonably long period of time (e.g., on the order of 18-24 hr). If flow through the fracture is controlled

by unsaturated conductivity in the surrounding soil media and assuming the conductivity is about 10 cm/s, -6

then the retention time in a fracture of 5 mm thickness could be on the order of a day or more depending on

the effective porosity in the iron. Thus, adequate retention time can be reasonably expected. As a treatment

barrier, the reactivity of the iron surface would need to exist for an extended period (e.g., 2-5 years or more).

Analysis of the micromorphology of the surface of the unused and used iron metal revealed some corrosion

of the iron surface after emplacement for ~10 months. The effect was limited however to only a fraction of

the available iron surface and this corrosion had no apparent effect on TCE degradation. Whether this would

hold true for longer emplacement periods is currently unknown.

With permanganate fractures, the dissolution of the solid permanganate particles will yield MnO ions that 4

-

will enter the soil solution and migrate away from the original location of emplacement by advection and/or

diffusion. This will yield an oxidizing zone that can treat TCE within the affected zone enveloping the

fracture as well as intercept and degrade mobile contaminants that enter the oxidation zone. This behavior

suggests that the permanganate fractures will be somewhat self-healing internally and between horizontally
Treatment Walls NATO/CCMS Pilot Study Phase III

81

(a) Media redox potential and pH above and below the horizontal

treatment fracture.

Figure 3. Features of permanganate horizontal barriers as

observed 10 months after initial emplacement (after Siegrist et al.,

1997).

(b) TCE degradation potential above and below the horizontal treatment

fracture as measured in batch tests using 3-5 g of media in 40 ml of

groundwater.

proximal fracture emplacements. With

regard to longevity, the permanganate

solids present in a 5 mm thick fracture

contains about 0.4 g KMnO per cm or 4

2

fracture horizontal area Based on

complete oxidation and a stoichiometric

dose of ~2.5 wt./wt., each cm of 2

fracture can treat ~0.16 g of TCE, which

is equivalent to roughly 16 l of percolate

with a concentration of 10 mg/l of TCE.

At a deep percolation flux of 1 cm/day,

this potential is equivalent to about 50

years of life. Realistically though, it is

anticipated that advective loss of oxidant

out of the treatment region or the oxidant

demand of natural organic matter will

markedly diminish this life. Based on

direct observation in this study, the

oxidation capacity within and around the

permanganate fracture was very high

even after 10-15 months of

emplacement. If alkenes, such as TCE,

were migrating downward in percolating

groundwater at reasonably high levels

(e.g., 10-100 mg/l), the capacity of the

fractures seemingly would be high

enough to intercept and treat this mobile

TCE.

The cost of a horizontal treatment barrier

comprised of reactive fractures such as

tested in this study were estimated.

When compared to standard hydraulic

fracturing and SVE, the major cost differences would be added costs for the reactive media (~$1.6/lb for

permanganate and ~$0.38/lb for iron metal) as compared to standard fracturing sand (~$0.10/lb) but lower

operational costs. The costs for media per fracture amounts to roughly $100 for sand, $1,000 for iron, and

$1,500 for permanganate. This increases the installation cost for the iron and the permanganate fractures.

However with lower resource consumption (e.g., power), less sampling and analysis (e.g., no off-gas),

reduced labor requirements, and no off-gas treatment costs, the operational costs would be substantially

lower for the horizontal barriers as compared to fracture-enhanced vapor extraction. For a 2.2 ha site and

a 5-m deep zone of contamination, the costs for emplacement of horizontal treatment barriers was estimated

to be in the range of $30/m . 

3

In summary, horizontal treatment barrier systems appear viable for TCE contaminated sites with silt and clay

deposits. Further work is necessary and appropriate, however, to fully develop this horizontal treatment

barrier approach and provide needed design and performance data for a range of full-scale applications. For

example, information is still needed on the depth and interval of emplacement as well as reactivity and

degradation capacity over time. The behavior of horizontal barrier systems under conditions of forced

advection also needs to be evaluated to understand the benefits/costs of recirculation approaches to in situ

treatment and source area mass reduction as opposed to more passive barriers for simple interception and

treatment.
Treatment Walls NATO/CCMS Pilot Study Phase III

82

4.0 References

American Petroleum Institute. 1995. Petroleum Contaminated Low Permeability Soil: Hydrocarbon

Distribution Processes, Exposure Pathways and In Situ Remediation Technologies. Health and

Environmental Sciences Dept. Publication No. 4631. Sept. 1995.

Case, T. 1997. Reactive Permanganate Grout for Horizontal Permeable Barriers and In Situ Treatment of

Groundwater. M.S. Thesis. Colorado School of Mines, Golden, CO. 

Case, T.L. and R.L. Siegrist. 1997. Oxidation Kinetics of Trichloroethylene in Groundwater by Permanganate under Varying Matrix Conditions. J. Water Environment Research. In review.

DOE. 1996. In Situ Remediation of DNAPL Compounds in Low Permeability Media: Transport/Fate, In

Situ Control Technologies, and Risk Reduction. Oak Ridge National Laboratory Report, ORNL/TM13305, for the U.S. Department of Energy. August, 1996.

Gates, D.D., R.L. Siegrist, S.R. Cline. 1995. Chemical Oxidation of Volatile and Semi-Volatile Organic

Compounds in Soil. Proc., Air and Waste Management Association Conf. June.

Murdoch, L., B. Slack, B. Siegrist, S. Vesper, and T. Meiggs. 1997. Hydraulic Fracturing Advances. Civil

Engineering. May 1997. pp. 10A-12A.

O’Hannesin, S.F. and R.W. Gillham. 1998. Long-term Performance of an In Situ “Iron Wall” for

Remediation of VOCs. Ground Water. 36(1):164-170.

Siegrist, R.L., O.R. West, J.S. Gierke, et al. 1995. In Situ Mixed Region Vapor Stripping of Low

Permeability Media. 2. Full Scale Field Experiments. Env. Sci. and Technol. 29(9):2198-2207.

Siegrist, R.L., K.S. Lowe, L.C. Murdoch, D.A. Pickering, T.L. Case. 1997. In Situ Oxidation by Fracture

Emplaced Reactive Solids. J. Environmental Engineering. In review.

U.S. EPA. 1992. Dense Nonaqeuous Phase Liquids—A Workshop Summary. EPA/600/R-92/030. Office

of Research and Development, Washington, D.C. 20460. 

U.S. EPA. 1993. Hydraulic Fracturing Technology—Technology Evaluation Report. EPA/540/R-93/505.

Office of Research and Development, Cincinnati, OH.

Discussion

Noting the rapid release rate for permanganate, Resat Apak asked whether permanganate sheets can be used

only in low-permeability aquifers. Siegrist acknowledged that his work involves the use of permanganate

grout in low-permeability zones; however, there is work being done with controlled releases of permanganate. With a slower release rate, perhaps the technology can be tailored to higher permeability zones.

Apak also whether Siegrist has observed clogging due to the formation of manganese hydroxide. Siegrist

noted that the product of the reaction between manganese and TCE is manganese oxide precipitate, which

ranges in size from 1-2 μm. The formation of the solids is somewhat related to the concentration of reactants.

Siegrist has observed limited formation of solids at the test site, but has not seen an accumulation at the

interface. Clogging has not been a problem with low-permeability zones, but may be a problem with higher

permeability zones.
Treatment Walls NATO/CCMS Pilot Study Phase III

83

Paul Bardos questioned whether the capacity of the permanganate is reduced due to oxidation of organic

matter in the soil. Siegrist agreed that permanganate is an oxidant that reacts with TCE and natural organic

matter in the soil. However, permanganate does not seem to be consumed as much by the organic matter.

Siegrist noted that residual permanganate remains in zones with organic material after 15 months; however,

the soil matrix should be assessed for the potential to consume the permanganate.

The cost difference between using permanganate and iron is not significant. The cost per fracture of using

sand alone is approximately US$100. The cost per fracture is US$1,000 for iron and US$1,500 for

permanganate. The retention time in the iron-filled fracture is 1-5 days, during which there is 30-40 percent

degradation. Siegrist noted that the cost estimate of US$30/m that he cited during his presentation applies 3

to both iron and permanganate. Both materials appear to be viable, although they have different

characteristics.
Treatment Walls NATO/CCMS Pilot Study Phase III

 UFZ – Centre for Environmental Research Leipzig-Halle, Department Industrial and Mining Landscapes, P.O. Box 2, 1

D - 04301 Leipzig, Germany 

 “Sanierungs-Forschung In Regional kontaminierten Aquiferen” – Remediation Research in Regionally Contaminated 2

Aquifers

84

In Situ Remediation Research in a Complexly Contaminated Aquifer:

The SAFIRA Test Site at Bitterfeld, Germany

H. Weiss , F.-D. Kopinke, P. Popp, and L. Wünsche 1

The experience of the last 20 years has shown that hydraulic soil and groundwater remediation schemes are

often very ineffective. This is especially true of large-scale pollution, where the source of contamination

either cannot be precisely located or is problematical to remove. Owing to the frequently low solubility of

hydrophobic organic substances (for example), the existence of a separate residual phase, the usually uneven

subterranean flow, and the sometimes slow return diffusion of pollutants that have penetrated the rock over

a period of years or even decades, solely hydraulic remediation can only achieve limited clean-up. The

consequences are very prolonged (and thus expensive) pump-and treat remedies, whose completion is

usually impossible to predict. 

Against the background of the now generally acknowledged problems of pump-and-treat activities, intensive

work has been carried out in recent years on developing low-price in situ remediation measures. The

currently most advanced remediation technology working in the flow is that of “in situ reaction/adsorption

walls” [1], which have already been tried out for simple pollutant mixtures (e.g., LCKW, BTEX, and PAK)

at a number of model sites. Nevertheless, there is still a great deal of development work to be done with

respect to the complex pollutant mixtures commonly encountered above all in the vicinity of abandoned

chemical plants.

As this problem is extraordinarily important in connection with the large-scale ecological remediation

schemes in eastern Germany, the SAFIRA project has been set up to examine and further develop the usage 2

of in situ reaction walls at a contaminated model field location.

In order to create the necessary basis for this model project, a preliminary study was carried out by UFZ in

cooperation with the Universities of Dresden, Halle, Leipzig, Stuttgart, and Tübingen. During the course of

this preliminary study, the hydrogeological and hydrogeochemical conditions at the planned field location

were investigated and various technologies for the design of the in situ pilot plant were examined and tested.

Choosing the Model Location

The Bitterfeld region was selected as the model location for investigations into developing powerful in situ

technologies for the remediation of complexly contaminated groundwater.

The soil and water environmental compartments in the Bitterfeld/Wolfen district have suffered sustained

damage as a result of over a century of lignite-mining and chemical industry. Whereas relevant soil pollution

is mainly confined to industrial locations (plant sites) and landfills, the persistent penetration of the

groundwater by pollutants has resulted in contamination attaining a regional scale. Consequently, an area

of about. 25 km with an estimated volume of some 200 million m is now partly highly polluted and must 2 3
Treatment Walls NATO/CCMS Pilot Study Phase III

85

Figure 1. Geological section (E–W) at the experimental site in the Bitterfeld area

be regarded as an independent source of contamination [2]. This pollution is characterized by the extensive

distribution of halogenized hydrocarbons, especially chlorinated aliphatics and chlorinated aromatics.

The development of the hydraulic conditions and thus regional pollutant transport is determined by the

draining techniques used for lignite-mining and its subsequent abandonment, as well as the imminent

flooding of Goitsche opencast mine. Since 1990, the hydraulic and conditions have been described and

evaluated in several studies (e.g., [2, 3, 4]; for further reading see [5]), so that there is an extraordinarily high

level of knowledge concerning the area in question. The containment and remediation concepts drafted since

1994 [2-4] feature both “active” and “passive” hydraulic containment measures to minimize groundwater

mobility. Successive measures to contain landfills (e.g., encapsulation) are also being discussed. “Active”

measures for groundwater remediation on a regional scale are not under consideration owing to the volumes

involved and the ad infinitum emission of pollutants from the landfills.

Given this background, Bitterfeld appears especially appropriate as the SAFIRA model location for

developing methods and technologies for minimally invasive “passive” decontamination techniques for

complexly polluted groundwater within a genuine scenario, as well as demonstrating their suitability in the

field. The results of the feasibility study for the experimental site are compiled in [6]. Figure 1 contains a

geological section of the area under investigation.

Pollution

Groundwater

Contamination of the groundwater with inorganic pollutants (e.g., heavy metals, arsenic, etc.) has proved

minor. The only noteworthy feature is the high levels of sulphate (up to 1,000 mg/l) and chloride (1,300

mg/l).

In order to characterize the organic pollution of the groundwater at the various water levels, several samples

were examined from different levels. The “main components” of organic contamination were found to be
Treatment Walls NATO/CCMS Pilot Study Phase III

86

chlorobenzene, 1,2-dichlorobenzene, 1,4-dichlorobenzene, benzene, trichloroethene, cis-1,2-dichloroethene

and trans-1,2-dichloroethene. Mass-spectrometry screening also detected other substances, albeit in very low

concentrations.

Although high organic pollution with halogenated hydrocarbons was confirmed in both the upper and lower

aquifers, contamination considerably varied between the two in terms of both quality and quantity. The lower

aquifer is dominated by aliphatic chlorinated hydrocarbons (trans-1,2-DCE, cis-1,2-DCE and TCE), while

the main component of the upper aquifer is chlorobenzene. During depth-oriented sampling, contamination

was found to display significant stratification: whereas the sample from a depth of 7-8 m only contains very

low amounts of pollutants, samples from depths between 16 and 20 m were above all found to contain high

levels of chlorobenzene (8-51 mg/l), as well as substantial concentrations of dichlorobenzenes (up to 1 mg/l).

Solid samples from the cores

In addition to the water samples, core material was also examined. The core samples from bore Saf Bit 2/96

were subjected to headspace-screening over the entire profile between 0 and 48.5 m. The relative sensitivities, which are proportional to the concentration of the substances in the sample, are summarized in Figure

2, showing clearly the influence of the lignite seam on the contaminant distribution.

One striking aspect is that the aliphatic halogen hydrocarbons and benzene chiefly occur in the strata

between 19.5 m and 24.5 m, while chlorobenzene and the dichlorobenzenes are found in higher concentrations in the strata between 12 m and 22.65 m. The main components are trans-1,2-dichloroethene, cis-1,2-

dichloroethene, chlorobenzene, trichloroethene, and benzene. Bromobenzenes and perchloroethene were also

identified by mass spectrometry.

Investigations into the adsorption and desorption of the pollutants in coal samples revealed that the sorption

capacity of the coal has not yet been exhausted. Hence the coal can act as both a pollutant sink (high

concentrations in the groundwater) and a pollutant source (lower concentrations of the pollutants in the

groundwater).

Laboratory Experiments on Pollutant Decomposition

Bioremediation

The basic requirements for deciding whether microbiological in situ remediation techniques can in principle

be used, selecting the most suitable technique and assessing the prospects of success are as follows:

C Determining the microbial density.

C Determining the pollutant breakdown capacity of the autochthonous microbial biocenoses under

in situ conditions.

C Examining ways of boosting the breakdown capacities of autochthonous microbial biocenoses

in situ using technically and financially feasible measures.

C Determining the microbial density.

C Determining the pollutant breakdown capacity of the autochthonous microbial biocenoses under

in situ conditions.
Treatment Walls NATO/CCMS Pilot Study Phase III

87

Figure 2. Contaminant profile, core 2/96

C Examining ways of boosting the breakdown capacities of autochthonous microbial biocenoses

in situ using technically and financially feasible measures.

In addition to the pollutant level, the aquifers at the location are characterized as microbial habitats by certain

combinations of microbe-ecologically relevant abiotic milieu factors. The quaternary aquifer has a lack of

oxygen. Sulphate and nitrate are present as potential electron acceptors for anaerobic processes. The pH

levels are around 7 and the temperatures about 13°C, only varying slightly along the depth profile.

The groundwater and aquifers are inhabited by bacteria down to the maximum drilling depth (50.5 m below

the surface). Nitrate, iron, and manganese reducing bacteria predominate in the autochthonous biocenoses.
time [d]

0 10 20 30 40

0

20

40

60

80

100

time [d]

0 10 20 30 40

Benzene

MCB

1,2-DCB

1,4-DCB

time [d]

0 10 20 30 40

0

20

40

60

80

100

time [d]

0 10 20 30 40

time [d]

0 10 20 30 40

A B

C D E

%

%

Treatment Walls NATO/CCMS Pilot Study Phase III

88

Figure 3. Decomposition of chlorobenzenes and benzene

by autochthonous bacteria groups in the boundary layer

of the quaternary aquifer and coal (19-23 m beneath the

surface)

Figure 3. a) Without the addition of substrates (A

aerobic; B-E anaerobic; B without additional

electron acceptors; C nitrate; D sulfate; E

hydrated iron oxide)

b) With the addition of substrates (lactate,

acetate, yeast extract, ammonium, and

phosphate); A-E

As expected, a lack of yeasts and mycelium-forming fungi was found in most habitats; a tiny abundance

(<5×10 g ) was only detected in the coal-bearing strata, individual horizons of the quaternary aquifers (6, 2 –1

16–17 and 19–21 m) and the tertiary aquifer at 48 m.

The pollutants found to be the main contaminants in the quaternary aquifer were selected for investigations

into microbiological breakdown. These were monochloro-benzene (MCB, 30 mg/l), 1,2-dichlorobenzene

(1,2-DCB, 10 mg/l) and 1,4-dichlorobenzene (1,4-DCB, 10 mg/l). Benzene (100 mg/l) was also included

in the investigation as the expected product of the chemically catalyzed dehalogenation of chlorobenzenes.

The three primary pollutants MCB, 1,2- and 1,4-DCB (as well as benzene) were removed from the autochthonous bacteriocenoses under simulated in situ conditions within technologically relevant periods of

times. However, the extent and speed of removal significantly differed depending on the pollutants’

chemical structure, the metabolism type of the ecophysiological groups involved, and the availability of

electron acceptors and additional carbon substrates (Figure 3).

Of the three chloroaromatics, 1,2-DCB easily proved to be the hardest compound to break down; in fact

under anaerobic conditions no decomposition whatsoever was established without the availability of

additional carbon substrates. After adding non-resident carbon substrates (a mixture of acetate and lactate),

the 1,2-DCB was however almost completely removed after just 20 days under nitrate, sulphate and iron

reducing conditions. When using these approaches, the other pollutants were no longer detectable either by

this time. 

Under aerobic conditions, 1,2-DCB was in all experiments only broken down very slowly (if at all). The

residual concentrations found after an incubation time of 40 days were usually 40 – 50 %, and in just one

case about 30% of the initial concentration - thus on the whole they were only slightly below the residual

concentration of 50 % defined as the evaluation limit for biological removal.
Treatment Walls NATO/CCMS Pilot Study Phase III

89

Assuming sufficient quantities of terminal electron acceptors and carbon substrates were available, the

pollutants were, somewhat surprisingly, broken down relatively quickly in the anaerobic environment under

nitrate-, sulphate-, and iron-reducing conditions.

The findings of the screening programme indicate that in principle the microbiological in situ remediation

of the contaminated aquifers under the anaerobic conditions prevailing at the location is possible. Furthermore, the in situ removal capacities can be increased by stimulating autochthonous anaerobic groups via the

addition of electron acceptors and usable carbon substrates.

Abiotic pollutant removal

Physical and chemical methods for pollutant reduction in the groundwater are grouped together within the

SAFIRA Project under the term abiotic techniques. Starting from the familiar process of dehalogenating

aliphatic chlorinated hydrocarbons with FeE walls, new methods were sought that above all enable the

removal of chloroaromatics and are suitable as in situ techniques. Initial approaches are seen in electrochemical techniques, sonochemical methods and bimetal systems. In addition, the usage of activated carbon

adsorption was tested, which could in particular be interesting as an in situ method if the residence times

were extended by microbial inhabitation. The goal of the preliminary study was to assess the possibilities

of such methods and to find new approaches. In the following, the findings of electrochemical removal are

presented by way of example; it should be pointed out that other approaches also yielded very promising

results.

A distinction was made between two different techniques for electrochemical dechlorination: systems using

palladium as a catalyst for hydrogen activation and those without palladium. The former represent the

current state of research, whereas the latter have not previously been described for chlorobenzene

degradation.

Pd was either precipitated electrochemically on the carbon cathode or mixed with the cathode material as

a carrier catalyst (e.g., 5% Pd on activated carbon). The main difference between the two methods is that

in the first case the Pd is on the electrode potential, whereas in the second it is not conductively linked to

the electrode.

In both methods, water electrolysis produces molecular hydrogen, which is activated at the surface of the

Pd. The concentration of chlorobenzene in the electrolytes (Bitterfeld groundwater) in these experiments

was 50–150 ppm. In many cases, chlorobenzene was completely removed. Figure 4 shows a typical course

of concentration over time for a mixture of mono-, di-, and trichlorobenzenes. The constancy of the chlorobenzene concentrations was measured up to 450 hours in order to record sorption and other losses. After

starting water electrolysis, the chlorobenzenes were mostly removed within about 50 hours. The regular

arrows indicate the renewed addition of the chlorobenzene mixture; the jagged arrows mark the beginning

of electrolysis. This sequence can be repeated several times without a drop in activity. The only reaction

products found were benzene and chloride (>90% of the converted chlorobenzenes). Similar results were

obtained when Pd on different carriers (activated carbon, soot) was freely suspended as a powder in the

groundwater.

Scaling Up

By now, the main site-specific investigations have been completed. The location has been found to be

suitable and the fundamental mechanisms for removing the “pollutant cocktail” are known. The investigations still underway are mainly designed to refine the data basis for hydraulic and matter transport modeling,

and to test and optimize additional removal techniques. Moreover, the technical principles for structural

planning are being worked out.
600 700 800

C h lo r o b e n z e n e

D ic h lo r o b e n z e n e

T r ic h lo r o b e n z e n e

1000000

2000000

3000000

100 200 300 400 500

0

0

E q u ilib r a t io n t im e ( v o lt a g e o f f ) A d d it io n o f s t o c k s o lu t io n

tim e [h ]

E le c t r o ly s is ( 5 V D C )

Treatment Walls NATO/CCMS Pilot Study Phase III

90

Figure 4. Reductive dechlorination of mono-, di- and trichlorobenzenes (c = 100 ppm MCB, 30 ppm DCB and 0

10 ppm TCB) on Pd (5 wt-% on 15 mg activated carbon) with hydrogen produced by electrolysis (cathode:

150 mg electrographite, U = 5 V, I = 2 mA, electrolyte: 150 ml groundwater).

The next stage is to scale up the experiments carried out so far in the laboratory. For this purpose a mobile

decontamination unit has been designed as a “window in the aquifer”. Groundwater from a depth of about

20 m is pumped into a storage tank without coming into contact with oxygen from the air. This polluted

water will then be used to charge five possible test columns with the physico-chemical conditions of the

aquifer being preserved (T = 13°C, pH = 7, eH = -250 mV, P = 2.5 mS/cm, AOX  100 mg/l).

The methods tested successfully in the laboratory and in the mobile decontamination unit have to prove their

chemical and hydrological long term stability and will be optimized in a pilot plant. Several vertical,

experimental columns (up to 3 m in diameter) will be installed to a depth of 20 m directly into the aquifer.

Numerous sampling and process controlling facilities and a variable design of the reaction columns will

enable the analyses of relevant chemical and hydraulic processes during operation and a competitive

technology development under real-world conditions. Operation of the pilot plant will start by the end of this

year. Subsequently an extension of the pilot plant for a horizontal groundwater flow is intended.

References

[1] Teutsch, G, Grathwohl, P., Schad, H. and Werner, P. (1996): In situ-Reaktionswände - ein neuer Ansatz

zur passiven Sanierung von Boden- und Grundwasserverunreinigungen.- Grundwasser 1/96, 12-20.

[2] Peter, H., Großmann, J. and Schulz-Terfloth, G. (1995): Rahmensanierungskonzept des Großprojektes

“Bitterfeld/Wolfen”.- In: LÜHR, H.-P. (Hrsg.): Grund-wassersanierung 1995.- IWS Schriftenreihe, 23:

123-138; Berlin, Erich Schmidt-Verlag

[3] Schulz-Terfloth, G. and Walkow, F. (1996): Maßnahmen zur Sanierung des Grundwassers unter

Berücksichtigung der wasserwirtschaftlichen und bergbaulichen Situation im Großprojekt BitterfeldWolfen.- In: LÜHR, H.-P. (Hrsg.): Grund-wassersanierung 1996.- IWS Schriftenreihe, 27: 307-320;

Berlin, Erich Schmidt-Verlag

[4] Landratsamt Bitterfeld / GFE GmbH: Grundwassermonitoring 1992-1996.
Treatment Walls NATO/CCMS Pilot Study Phase III

91

[5] Großmann, J. and Lühr, H.-P. (1994): Sanierungsrahmenkonzept für Großprojekt Bitterfeld-Wolfen.

[6] Weiß, H., Teutsch, G. and Daus, B. (Hrsg.) (1998): Sanierungsforschung in regional kontaminierten

Aquiferen (SAFIRA) - Bericht zur Machbarkeitstudie für den Modellstandort Bitterfeld.- UFZ-Bericht

27/1997, ISSN 0948 94252; Januar 1998; Leipzig

Discussion

Kahraman Ünlü asked whether the removal of iron chloride is a concern. Weiss noted that lowering chloride

concentrations is not feasible due to the presence of underground brine.

Weiss was asked if he considered using iron to produce hydrogen because it is cheaper than producing it

electrolytically. Weiss indicated the oxygen that is produced in the electrolytical reaction can be used for

aerobic reactions.

One of Weiss’s photographs showed gas bubbles being produced at the water surface during the reaction,

which were identified as methane and nitrogen.
Treatment Walls NATO/CCMS Pilot Study Phase III

92

Summary and Conclusions

Harald Burmeier

Construction of Treatment Walls

The costs of pilot studies and field studies should be considered when computing the costs of implementing

a technology. The costs presented for the funnel-and-gate technology only reflect the materials (iron) and

construction. Field demonstrations are not conducted very frequently anymore; in most cases, laboratory

studies suffice. At approximately 90 percent of the North American sites for which treatment walls, TCE

or PCB is the primary contaminant; with few exceptions, the half-lives fall within a narrow range (30

minutes to 1 hour). Therefore, it becomes difficult to justify field tests as cost-effective.

Eberhard Beitinger recently performed field tests with two columns of activated carbon at a site with several

types of contaminants. The columns were installed on site to model actual conditions. The results of the

column tests were very encouraging; the performance time of the wall would be higher than had been

expected. The minimum time designed for the wall was 30 years for the first removal of reactive materials.

A 50-m continuous wall will be installed with impermeable “wings.”

Although residence times to degrade chlorinated solvents are generally known, geological and hydrogeological characterization is needed for any treatment approach. Furthermore, information on the plume’s

location and where it is moving is also necessary. If this information is well understood, design costs can

be lowered.

Harry Whittaker doubted that chromium can be removed permanently with an iron treatment wall because

iron becomes very finely divided ferric oxide in the environment, especially in wet soils. He suggested that

there is a continuous long-term cycle of Fe , Fe , FeE, and Cr and Cr . Once all of the iron is converted 2+ 3+ 3+ 6+

to Fe , the chromium that was retained by the treatment wall will be released. Wolfgang Wüst added that 3+

he has seen this occur in column tests.

Bob Puls said that the state environmental agency in North Carolina had to be convinced that chromate

would not be released eventually from the permeable reactive barrier (PRB) at Elizabeth City. Originally,

natural attenuation of the chromate was examined. There was a natural reduction of chromate observed

(primarily due to the high organic carbon content of the surface soils), and the chromate was immobilized

in a chromium and iron hydroxide phase. Researchers had to prove to the regulatory authority that the iron

would not reoxidize the Cr to Cr . The only oxidant in natural soil strong enough to oxidize chromium 3+ 6+

is manganese oxide. As a result, they analyzed the manganese oxide content of the soil and groundwater to

prove that whatever immobilization method was occurring would not reverse itself. The PRB was not

considered a permanent remedy for the chromate plume, and they are looking into eliminating the chromate

source.

Harald Burmeier said that long-term aspects of installing treatment walls must be considered during the

feasibility study. For example, the construction of the wall should be approached so that the reactive media

can be removed and replaced as necessary. Although it may be relatively inexpensive to construct a

treatment wall, the need to remove it 30 years in the future must be considered in cost estimates.

Paul Bardos asked if there is any true distinction between permeable reactive barriers and treatment walls.

Burmeier explained that permeable reactive barriers are a type of treatment wall that uses a material that is

permeable and reactive to treat the groundwater. Treatment walls can also be partially permeable and treat
Treatment Walls NATO/CCMS Pilot Study Phase III

93

groundwater using adsorptive, biological, or electrochemical processes. In addition to treatment wall

construction, treatment barriers can be constructed as piles, drainage trenches, and funnel and gate systems.

There was additional discussion on the issue of patents for treatment walls. To the extent that EnviroMetal

Technologies has a worldwide patent on the emplacement of iron in the ground for remediation of

chlorinated organic solvents, there may be cases where researchers are not adhering to the patent restrictions.

Hermann Schad indicated that the funnel-and-gate patent applies to situations where groundwater flow is

diverted toward the treatment wall. This design is protected in Germany and several other countries in the

European Union. Schad noted that funnel-and-gate systems are advantageous when intending to use a

material to react with or absorb a contaminant. The continuous walls, however, have hydraulic advantages.

In response to a question regarding treatment walls that do not work as designed, Burmeier acknowledged

that the examples presented in the session emphasized successes rather than failures. However, he

emphasized that reactive materials must be tested prior to emplacement. He added that when building a wall

over 10 to 20 m deep, technologies must be matched to the permeability of the subsurface as well. Gillham

was aware of at least three cases in the United States where regulators required a contingency for the

installation of a reactive barrier. He added that in any case, an alternative treatment method is almost always

proposed along with the reactive barrier.

Walt Kovalick emphasized that the cost savings of using PRBs instead of pump-and-treat systems is significant. He suggested that 10 to 20 percent of the pump-and-treat systems currently active in the United States

could benefit from the installation of PRBs. The issue of contingencies was insignificant from a cost

standpoint because the cost of a typical wall (US$400 thousand) is significantly less than a pump-and-treat

system (US$12-14 million). Therefore, the cost of research and development is also justified.

Reactive Materials, Full-Scale Projects, and General Conclusions

Harald Burmeier summarized the topics discussed and the questions on treatment walls that still remain to

be answered. He pointed out that the Special Session presentations have examined the reaction, absorption,

and degradation processes of treatment walls, but noted that combinations of processes are also being

considered.

Several issues need closer examination, such as how to estimate and influence the long-term performance

of treatment wall; determination of by-products of organic contaminant degradation; the cost of materials

(e.g., the range in costs and corresponding effectiveness of zero-valent irons); calculation of mass balances;

the determination of what reactive substances can be placed in the groundwater without risk; and how to

influence biological processes in combination with treatment walls.

The presentations showed that zero-valent iron can be used to degrade chlorinated solvents and adsorb

uranium, but the nature and toxicity of the adsorbed uranium still must be determined. Clearer explanations

are needed from scientists on how to evaluate degradation and sorption processes. Optimization of reactive

materials using bench-scale tests have frequently been replaced by laboratory tests followed by trial-anderror field testing. Further research on bioscreens is needed to determine long-term effects on remediation

as well as possible limiting factors such as bioclogging, redox limitations, and concentrations or

combinations of contaminants. There are questions beyond technological solution, that cross into regulatory,

legal, and public policy, such as what concentration of a contaminant is harmless? Can reactive materials

be emplaced in the groundwater, and at what concentrations? What is the long-term implication of in situ

chemical engineering, including any production of toxic by-products? 
Treatment Walls NATO/CCMS Pilot Study Phase III

94

From the standpoint of constructing treatment walls, good definitions of terms such as “permeable reactive

barriers” and “treatment walls” must be developed. Although it is now possible to excavate and install deep

walls, the cost effectiveness of such constructions should be evaluated in order to access the market. 

In conclusion, treatment wall technology has been proven to work; however, there are some limitations to

its use. Several questions still must be answered before their use can become widespread.
Treatment Walls NATO/CCMS Pilot Study Phase III

95

ABOUT THE AUTHORS

Harald Burmeier is a civil engineer with 20 years of experience in waste and hazardous waste management,

contaminated site investigations, feasibility studies, and health and safety issues. A professor of engineering

on the faculty of Applied Studies and Research of the University of Northeast Lower Saxony, Prof. Burmeier

maintains an active consulting practice throughout Germany on project management for waste and hazardous

waste remediation, issues related to worker health and safety, and emissions control. He is the author of a

number of books and articles on hazardous waste investigation, management, and remediation. Prior to

assuming his present position last year, Prof. Burmeier held a series of senior technical and management

positions in the German construction industry and trade associations. Prof. Burmeier holds a M.S. in Civil

Engineering from the Technical University of Hannover, Germany.

Gérard Evers is a senior engineer and head of the Department of Technique and Development at Soletanche

Bachy (Nanterre, France), the world’s largest specialized contractor in foundation techniques and

underground construction. Mr. Evers holds a diploma in civil engineering from the Technical High School

of The Netherlands.

Robert Gillham has been a professor in the Department of Earth Sciences at the University of Waterloo

since the early 1970s. He was Director of the Waterloo Centre for Groundwater Research from 1987 to 1992,

was Chair of the Earth Sciences Department from 1993 to 1997, and currently holds the NSERC/Motorola/

ETI Industrial Research Chair in Groundwater Remediation. In 1997 Dr. Gillham was elected a Fellow of

the Royal Society of Canada. Dr. Gillham’s primary research areas concern contaminant transport in

groundwater systems and groundwater remediation.

Liyuan Liang received her B.S. in civil engineering from Northeastern University and her M.S. and Ph.D.

in Environmental Engineering from the California Institute of Technology. Dr. Liang is Senior Lecturer in

the Department of Earth Sciences, University of Wales (UK), where she teaches aquatic geochemistry and

environmental remediation technologies, conducts research on physico-chemical processes, and directs the

Virtual Institute for Environmental Research. Prior to her current appointment, she was at Oak Ridge

National Laboratory (USA), where she managed a group of researchers investigating in situ phsico-chemical

treatment technologies for groundwater remediation and on the mechanistic understanding of geochemical

processes in the environment. She also conducted research on colloidal transport in groundwater, the kinetics

of reductive dechloriantion of organic solvents using zero-valent metals, and on the peroxidation of organic

contaminated soils. While at Oak Ridge, she was a member of the Steering Committee of the U.S. Permeable

Reactive Barriers Action Team. Her primary research interests include in situ treatment technologies for

mixed-wastes (organic solvents and radionuclides); laboratory and field research on particle dynamics in

aquatic environments (including coagulation, dispersion, and transport of colloidal particles); oxidative

precipitation and mineral dissolution; and fate of metals and organic contaminants in the environment. Dr.

Liang has published extensively in refereed journals. 

Stephan Jefferis (M.A., M.Eng., M.Sc., Ph.D., C.Eng., M.I.C.E., C.Geo.l, F.G.S.) has qualifications in

natural sciences, chemical engineering, civil engineering, and law. He spent 20 years, with the University

of London (UK) where his research group developed a class of cement-bentonite cut-off materials that has

been used at many contaminated sites in Europe. Dr. Jefferis is presently an associate of Golder Associates,

where his role is world-wide and typically involves the investigation and resolution of unusual geotechnical,

geoenvironmental and materials problems often associated with aggressive chemical and microbiological

processes in the ground. He is also a visiting professor in the Department of Civil Engineering, Imperial

College, London.
Treatment Walls NATO/CCMS Pilot Study Phase III

96

Robert W. Puls has a B.S. in Soil Science and Natural Resources from the University of Wisconsin, a

Masters degree in Forest Resources from the University of Washington, and a Ph.D. degree in Soil and

Water Science from the University of Arizona. He has worked as a Research Soil Scientist at the R.S. Kerr

Environmental Research Center, of the National Risk Management Research Laboratory of U.S. EPA in

Ada, Oklahoma since 1987. Prior to that he worked for DOE in Richland, Washington on the High Level

Nuclear Waste Repository Research Program. He has more than 19 years experience working in the

environmental science field. Dr. Puls is the current Co-Chair of the Permeable Reactive Barrier Action

Team, one of seven Action Teams comprising EPA’s Remedial Technology Development Forum (RTDF).

He has been actively involved in research related to permeable reactive barriers since 1991. His recent

publications have covered a range of topics including groundwater sampling, colloidal transport in

groundwater, organic-metal-mineral interactions, the development and evaluation of in situ remediation

techniques for soils and groundwater, and metal and metalloid sorption-desorption reactions governing

subsurface contaminant transport and transformation processes. He has authored more than 50 publications

on the above topics and given numerous presentations at national and international scientific meetings. Dr.

Puls has served on several scientific advisory committees for USEPA, the USGS, DOE and the National

Research Council.

Huub H. M. Rijnaarts earned his M.Sc. degree in soil chemistry and soil microbiology and his Ph.D. from

the University of Wageningen Agricultural University (The Netherlands), where he studied bacterial

transport in porous media and biodegradation of hazardous organic compounds. Since 1994, Dr. Rijnaarts

has been a Senior Researcher with The Netherlands Organization for Applied Scientific Research (TNO)

Institute of Environmental Sciences, Energy Research and Process Innovation, where he has been involved

in developing new in situ soil and groundwater bioremediation techniques, including biological barriers, and

natural attenuation.

Hermann Schad holds an M.S. and Ph.D. in geology from the University of Oregon (USA) and the

Unviersity of Tübingen (Germany), respectively. His Ph.D. thesis was on the “Variabliity of Hydraulic

Parameters in Non-Uniform Porous Media: Experiments and Stochastic Modelling at Different Scales.” Dr.

Schad presently is the Waste Management Director of I.M.E.S. (Innovative Mess-, Erkundunes- und

Sanierungstechnologien GmbH) in Wangen, Germany.

Robert L. Siegrist earned his B.S. and M.S. in Civil Engineering and his Ph.D. in Environmental

Engineering at the University of Wisconsin. During 20 years of experience, he has held research and

teaching appointments with the Colorado School of Mines, Oak Ridge National Laboratory (ORNL), the

University of Wisconsin, and the Agricultural University of Norway. Since January 1995, Dr. Siegrist has

been affiliated with Colorado School of Mines (USA) as a research associate professor in the Environmental

Science and Engineering Division while also holding an adjunct faculty participant position as a senior staff

member in the Environmental Sciences Division at ORNL. His research has focused on characterization,

assessment, and in situ remediation technologies for contaminated land. The technology related work has

included in situ treatment processes (chemical oxidation, redox reactive barriers, bioremediation) as well

as subsurface manipulation methods (vertical and horizontal recovery and recirculation wells, lance

permeation, deep soil mixing, hydraulic fracturing). In related work, Dr. Siegrist has continued research into

the hydrodynamic and purification processes impacting land treatment and disposal of wastes. Since 1990,

Dr. Siegrist has been leading a program of research concerning in situ chemical ozidation involving peroxide

and permanganate systems including fundamental studies of oxidation reaction chemistry and kinetics,

contaminant mass transfer, oxidant delivery systems, and field evaluations through pilot- and full-scale

technology demonstrations. His research has been sponsored by the U.S. Department of Energy, U.S. EPA,

the U.S. Department of Defense, the National Science Foundation, and private industry. He has published

his results in over 35 refereed articles and more than 100 conference proceedings and reports. He is an active

member of several national societies and currently servces on national committees, including the

Groundwater, Hazardous Waste, and McKee Medal Committees of the Water Environment Federation. he
Treatment Walls NATO/CCMS Pilot Study Phase III

97

is a registered professional engineer and has served as an advisor and technical expert for state and federal

agencies in the United States, Canada, Norway, and Sweden. He is currently a Fellow with NATO/CCMS.

Timothy M. Vogel received his B.S. degrees in biological oceanography and geology and his M.S. in

environmental engineering from the University of Washington (USA), and his Ph.D. in environmental

engineering (microbiology) from Stanford University (USA). Following his graduate studies, Dr. Vogel

taught at Michigan State University and the University of Michigan. Between 1992 and 1997, he held

positions of research and management with Rhône-Poulenc Industrialisation and, since 1998, he has been

the research and development project manager for Rhodia Eco Services in Meyzieu, France. His current

research activities include the acquisition, evaluation, and development of novel technologies for the

remediation of polluted soil and groundwater. Reactive barriers are among his most intense research

subjects, based on his previous scientific research involving the development of reductive dechlorination

systems—such as vitamin B12—and on the biodegradation of petroleum hydrocarbons.

Holger Weiss is Head of the Department of Industrial and Mining Landscapes at the Center of

Environmental Research, Leipzig-Halle (Germany), where he coordinates interdisciplinary and international

research on contaminated and degraded landscapes, minimization of environmental impacts of mineral

exploitation and development, landscape reclamation, and ecosystem stability in industrial and mining areas.

He is conducting research on in situ technologies for complexly contaminanted aquifers (active and passive;

physico-chemical; microbiological), on the environmental significance and human and ecological affects

from industrial residues, and on regional emissions and effects of choro-organic contaminants. Dr. Weiss

holds a Ph.D. in geology from the Technical University of Clausthal, Lower Saxony (Germany), and he has

published extensively in the peer-reviewed literature in the United States and Europe.

Wolfgang Wüst was born in Nördlingen (Bavaria, Germany). After high school he got a scholarship from

the state Bavaria and decided to study Geoökologie (Environmental Sciences) at the University of Bayreuth

and Lancaster (UK) with the focus on environmental chemistry and modeling of pollutant transport and

degradation. In 1990 he received his diploma on the transport of heavy metals from the street into sewer

systems in the working group of Professor R. Herrmann. He continued his training in the field of

hydrogeology and groundwater hydraulics at the chair of groudwater hydraulics (Prof. Kobus) and was

involved in different research projects related to groundwater remediation in the VEGAS facility at the

University of Stuttgart. Since three years he is focussing on the chemical processes defining the scope and

the limitations of permeable reactive iron walls (working group of Professor Dahmke) and will soon finish

his Ph.D. on this topic.
Treatment Walls NATO/CCMS Pilot Study Phase III

98

NATIONAL CONTACTS
Treatment Walls NATO/CCMS Pilot Study Phase III

99

Directors

Stephen C. James (Co-Director) Walter W. Kovalick, Jr. (Co-Director)

National Risk Management Research Laboratory Technology Innovation Office

U.S. Environmental Protection Agency U.S. Environmental Protection Agency

26 Martin Luther King Drive 401 M Street, SW (5102G)

Cincinnati, Ohio 45268 Washington, DC 20460

United States United States

tel: 513-569-7877 tel: 703-603-9910

fax: 513-569-7680 fax: 703-603-9135

e-mail: james.steve@epamail.epa.gov e-mail: kovalick.walter@epamail.epa.gov

Co-Pilot Directors

Volker Franzius H. Johan van Veen

Umweltbundesamt The Netherlands Integrated Soil Research

Bismarckplatz 1 Programme

D-14193 Berlin P.O. Box 37

Germany

tel: 49/30-8903-2496

fax: 49/30-8903-2285 or -2103 tel: 31/317-484-170

NL-6700 AA Wageningen

The Netherlands

fax: 31/317-485-051

e-mail: anneke.v.d.heuvel@spbo.beng.wau.nl

Country Representatives

Nora Auer Harry Whittaker

Federal Ministry of Environment, Youth and Emergencies Engineering Division

Family Affairs Environment Canada

Dept. III/3 3439 River Road

Stubenbastei 5 Ottawa, Ontario, K1A OH3

A-1010 Vienna

Austria

tel: 43/1-515-22-3449 fax: 613/991-1673

fax: 43/1-513-1679-1008 e-mail: harry.whittaker@etc.ec.gc.ca

e-mail: Nora.Auer@bmu.gv.at

Jacqueline Miller Aquatest a.s.

Brussels University Geologicka 4

Avenue Jeanne 44 152 00 Prague 5

1050 Brussels

Belgium

tel: 32/2-650-3183 fax: 420/2-581-77-58

fax: 32/2-650-3189 e-mail: aquatest@aquatest.cz

e-mail: jmiller@resulb.ulb.ac.be

Canada

tel: 613/991-1841

Jan Švoma

Czech Republic

tel: 420/2-581-83-80
Treatment Walls NATO/CCMS Pilot Study Phase III

100

Inge-Marie Skovgård Matthew Crowe

Contaminated Land Division Environmental Management and Planning

Danish Environmental Protection Agency Division

29 Strandgade Environmental Protection Agency

DK-1401 Copenhagen K P.O. Box 3000

Denmark

tel: 45/3-266-0100 - direct 45/32660397 County Wexford

fax: 45/3-296-1656

e-mail: ims@mst.dk tel: +353 53 60600

Ari Seppänen e-mail: m.crowe@epa.ie

Ministry of Environment

P.O. Box 399 Takeshi Nishio

00121 Helsinki Soil and Agricultural Chemicals Division

Finland

tel: +358/9-199-197-15 Japan Environment Agency

fax: +358/9-199-196-30 1-2-2, Kasumigaseki, Chiyoda-Ku

René Goubier

Polluted Sites Team tel: +81-3-3580-3173

ADEME fax: +81/3-3593-1438

B.P. 406 e-mail: takeshi_nishio@eanet.go.jp

49004 Angers Cedex 01

France

tel: 33/241-204-120 Ministry for the Environment

fax: 33/241-872-350 84 Boullcott Street

Antonios Kontopoulos Wellington *

National Technical University of Athens

GR-157 80 Zografos tel: 64/4-917-4000

Athens fax: 64/4-917-7523

Greece

Pál Varga Bjørn Bjørnstad

National Authority for the Environment Norwegian Pollution Control Authority

Fö u.44 P.O. Box 8100 Dep

H-1011 Budapest N-0032 Oslo

Hungary Norway

tel: 36/1-457-3530 tel: 47/22-257-3664

fax: 36/1-201-4282 fax: 47/22-267-6706

e-mail: vargap@kik.ktm.hu e-mail: bjorn.bjornstad@sftospost.md.dep.

Johnstown Castle Estate

Ireland

fax: +353 53 60699

Environment Agency, Water Quality Bureau

Tokyo 100

Japan

Raymond Salter

P.O. Box 10362

New Zealand

e-mail: rs@mfe.govt.nz

telemax.no

Ewa Marchwinska

Institute for Ecology of Industrial Areas 

6 Kossutha Street

40-833 Katowice

Poland

tel: 48/32 -1546-031

fax.: 48/32 -1541-717

e-mail: ietu@ietu.katowice.pl

*Due to the death of Prof. Kontopoulos,

communications with the Greek delegation to the Pilot

Study may be directed to: 

Manolis Papadopoulos, tel: +30-1-772 2219; fax:

+30-1-772 2218, e-mail papadop@metal.ntua.gr
Treatment Walls NATO/CCMS Pilot Study Phase III

101

Marco Estrela Resat Apak

Instituto de Soldadura e Qualidade Istanbul University

Centro de Tecnologias Ambientais Avcilar Campus, Avcilar 34850

Estrada Nacional 249–Km 3–Leiao (Tagus Park) Istanbul

Apartado 119 - 2781 Oeiras Codex

Portugal

tel: +351/1-422-8100 fax: 90/212-5911-997

fax: +351/1-422-8129 e-mail: rapak@istanbul.edu.tr

e-mail: maestrela@isq.pt

Branko Druzina Depratment of Environmental Engineering

Institute of Public Health Middle East Technical University

Trubarjeva 2-Post Box 260 Inönü Bulvari

6100 Ljubljana 06531 Ankara

Slovenia Turkey

tel: 386/61-313-276 tel: 90-312-210-1000

fax: 386/61-323-955 fax: 90-312-210-1260

e-mail: branko.druzina@ivz.sigov.mail.si e-mail: kunlu@rorqual.cc.metu.edu.tr

Ingrid Hasselsten Ian D. Martin

Swedish Environmental Protection Agency Environment Agency

Blekholmsterrassen 36 Olton Court

S-106 48 Stockholm 10 Warwick Road

Sweden

tel: 46/8-698-1179

fax: 46/8-698-1222 tel: 44/121-711-2324

e-mail: inh@environ.se fax: 44/121-711-5830

Bernhard Hammer

BUWAL

Federal Department of the Interior

3003 Bern

Switzerland

tel: 41/31-322-9307

fax: 41/31-382-1546

Turkey

tel: 90/212-5911-998

Kahraman Ünlü

Olton, West Midlands

United Kingdom

e-mail: ianmartin@environment-agency.gov.uk
Treatment Walls NATO/CCMS Pilot Study Phase III

102

PARTICIPANTS

Phase III Pilot Study Meeting

Vienna, Austria

February 23-27, 1998
Treatment Walls NATO/CCMS Pilot Study Phase III

103

Resat Apak Paul M. Beam

Istanbul University U.S. Department of Energy

Avcilar Campus, Avcilar 34850 19901 Germantown Road

Istanbul Germantown, MD 20874-1290

Turkey United States

tel: 90/212-5911-998 tel: 301-903-8133

fax: 90/212-5911-997 fax: 301-903-3877

e-mail: rapak@istanbul.edu.tr e-mail: paul.beam@em.doe.gov

Nora Auer Eberhard Beitinger

Federal Ministry of Environment, Youth and WCI Umwelttechnik GmbH

Family Affairs Sophie Charlotten - Str.33

Dept. III/3 14059 Berlin

Stubenbastei 5 Germany

A-1010 Vienna tel: 49/30-3260-9481

Austria fax: 49/30-321-9472

tel: 43/1-515-22-3449

fax: 43/1-513-1679-1008 Bjørn Bjørnstad

e-mail: Nora.Auer@bmu.gv.at Norwegian Pollution Control Authority

Erik Bäcklund N-0032 Oslo

Eko Tec Norway

Näsuddsvägen 10 - Box 34 tel: 47/22-257-3664

932 21 Skelleftehamn fax: 47/22-267-6706

Sweden e-mail: bjorn.bjornstad@sftospost.md.dep.

tel: 46/910-333-66 telemax.no

fax: 46/910-333-75

Paul Bardos University of applied Studies and Research

R3 Environmental Technologies Ltd Herbert-Meyerstrasse 7

P.O. Box 58 29556 Suderburg

Ware- Hertfordshire SG12 9UJ Germany

United Kingdom tel: 49/5103-2000

tel: 44/1920-484-571 fax: 49/5103-7863

fax: 44/1920-485-607 e-mail: h.burmeier@t-online.de

e-mail: p-bardos@r3-bardos.demon.co.uk

N. Jay Bassin Environmental Management Support, Inc.

Environmental Management Support, Inc. 8601 Georgia Avenue, Suite 500

8601 Georgia Avenue, Suite 500 Silver Spring, Maryland 20910

Silver Spring, Maryland 20910 United States

United States tel: 301-589-5318

tel: 301-589-5318 fax: 301-589-8487

fax: 301-589-8487 e-mail: ddopkin@emsus.com

e-mail: jbassin@emsus.com

P.O. Box 8100 Dep

Harald Burmeier

Diane Dopkin
Treatment Walls NATO/CCMS Pilot Study Phase III

104

Branko Druzina Inger Asp Fuglsang

Institute of Public Health Contaminated Land Division

Trubarjeva 2-Post Box 260 Danish Environmental Protection Agency

6100 Ljubljana 29 Strandgade

Slovenia DK-1401 Copenhagen K

tel: 386/61-313-276 Denmark

fax: 386/61-323-955 tel: 45/32-66-01-00

e-mail: branko.druzina@ivz.sigov.mail.si fax: 45/32-66-04-79

Erol Erçag

Istanbul University Robert Gillham

Dept. of Chemistry University of Waterloo

Avcilar Campus, Avcilar 34850 Department of Earth Sciences

Istanbul Waterloo, Ontario N2L 3G1

Turkey Canada

tel: 90/212-5911-998 tel: 519-888-4658

fax: 90/212-5911-997 fax: 519-746-7484

Marco Estrela

Instituto de Soldadura e Qualidade René Goubier

Centro de Tecnologias Ambientais Polluted Sites Team

Estrada Nacional 249–Km 3–Leiao (Tagus Park) ADEME

Apartado 119 - 2781 Oeiras Codex B.P. 406

Portugal 49004 Angers Cedex 01

tel: +351/1-422-8100 France

fax: +351/1-422-8129 tel: 33/241-204-120

e-mail: maestrela@isq.pt fax: 33/241-872-350

Gerard Evers Iliana Halikia

Soletanche National Technical University of Athens 

6 Rue De Wattford GR-157 80 Zografos

F92000 Nanterre Athens

France Greece

tel: 33/14-7764-262 tel: 30/1-722-2167

fax: 33/14-9069-734 fax: 30/1-722-2168

e-mail: gerard.evers@soletanche-bachy.com e-mail: labmet@metal.ntua.gr

Volker Franzius Bernhard Hammer

Umweltbundesamt BUWAL

Bismarckplatz 1 Federal Department of the Interior

D-14193 Berlin 3003 Bern

Germany Switzerland

tel: 49/30-8903-2496 tel: 41/31-322-9307

fax: 49/30-8903-2285 or -2103 fax: 41/31-382-1546

e-mail: iaf@mst.dk

e-mail: rwgillha@sciborg.uwaterloo.ca
Treatment Walls NATO/CCMS Pilot Study Phase III

105

Catherine Harvey Harald Kasamas

Environment Agency CARACAS - European Union

Steel House Breitenfurterstr. 97

11 Tothill Street A-1120 Vienna

London Austria

United Kingdom tel: 43/1-804 93 192

tel: 44/171-664-6793 fax: 43/1-804 93 194

fax: 44-171-664-6795 e-mail: 101355.1520@compuserve.com

e-mail: diane.williamson@environmentagency.gov.uk (office e-mail) Vladimír Kinkor

Ingrid Hasselsten P.O. Box 47

Swedish Environmental Protection Agency Bezecká 79

Blekholmsterrassen 36 169 00 Prague 6

S-106 48 Stockholm Czech Republic

Sweden tel: 420/602-347-679

tel: 46/8-698-1179 fax: 420/602-5721-1255

fax: 46/8-698-1222

e-mail: inh@environ.se Antonios Kontopoulos

Christian Holzer Athens

Department of Waste Treatment and Remediation Greece

of Abandoned Sites

Federal Ministry of Environment, Youth, and Walter W. Kovalick, Jr.

Family Affairs (Dept. III/3) Technology Innovation Office

Stubenbastei 5 U.S. Environmental Protection Agency

A-1010 Vienna 401 M Street, SW (5102G)

Austria Washington, DC 20460

tel: 43/1-515 22-3429 United States

fax: 43/1- 513 16 79 - 1127 tel: 703-603-9910

e-mail: christian.holzer@bmu.gv.at fax: 703-603-9135

Stephen C. James

National Risk Management Research Laboratory Tomáš Lederer

U.S. Environmental Protection Agency Aquatest a.s.

26 Martin Luther King Drive Geologika 4

Cincinnati, Ohio 45268 152 00 Prague

United States Czech Republic

tel: 513-569-7877 tel: 420/2-581-8995

fax: 513-569-7680 fax: 420/2-581-8175

e-mail: james.steve@epamail.epa.gov e-mail: lederer@aquatest.cz

Stephan Jefferis Liyuan Liang

Golder Associates (UK) Ltd. Department of Earth Sciences

54-70 Moorbridge Road University of Wales, Cardiff

Maidenhead, Berkshire P.O. Box 914

SL6 8BN England Cardiff GF1 34E

United Kingdom United Kingdom

tel: 44/1628-771-731 tel: 44/1-222-874-579

fax: 44/1628-770-699 fax: 44/1-222-874-326

e-mail: sjefferis@golder.com e-mail: liyuan@cardiff.ac.uk

SEPA s.r.o.

National Technical University of Athens

e-mail: kovalick.walter@epamail.epa.gov
Treatment Walls NATO/CCMS Pilot Study Phase III

106

Ian D. Martin Robert Puls

Environment Agency U.S. Environmental Protection Agency

Olton Court 919 Kerr Research Drive

10 Warwick Road P.O. Box 1198

Olton, West Midlands Ada, Oklahoma 74820

United Kingdom United States

tel: 44/121-711-2324 tel: 580-436-8543

fax: 44/121-711-5830 fax: 580-436-8703

e-mail: ianmartin@environment-agency.gov.uk e-mail: puls.robert@epamail.epa.gov

Igor Marvan H.H.M. Rijnaarts

Grace Dearborn Inc. TNO/MEP

3451 Erindale Station Road P.O. Box 342

P.O. Box 3060, Station A 7300 AH Apeldoorn

Mississauga, Ontario L5A 4B6 The Netherlands

Canada tel: 31/55-5493-380

tel: 905-272-7435 fax: 31/55-5493-410

fax: 905-272-7456 e-mail: h.h.m.rijnaarts@mep.tno.nl

Jacqueline Miller Hermann Schad

Brussels University IMES GmbH

Avenue Jeanne 44 Kocherhof 4

1050 Brussels 88239 Wangen

Belgium Germany

tel: 32/2-650-3183 tel: 49/7528-971-30

fax: 32/2-650-3189 fax: 49/7528-97131

e-mail: jmiller@resulb.ulb.ac.be e-mail: hermann.schad.imes@t-online.de

Walter Mondt Mathias Schluep

Ecorem n.v. BMG Engineering AG

Zwartzustersvest 22 Ifangstrasse 11

B-2800 Mechelen 8952 Schlieren

Belgium Switzerland

tel: 32-15-21 17 35 tel: 41/1-730-6622

fax: 32-15-21 65 98 fax: 41/1-730-6622

e-mail: Ecorem@glo.be

Carlos de Miguel Perales Eberhard-Karls-Universität Tübingen

ICADE Geologisches Institut

Alberto Aguilera, 23 Sigwarstr. 10

28015 Madrid 72076 Tübingen

Spain Germany

tel: 34/1-586-0455 tel: 49/7071-29-75041

fax: 34/1-586-0402 fax: 49/7071-5059

Christoph Schüth

e-mail: christoph.schueth@uni-tuebingen.de
Treatment Walls NATO/CCMS Pilot Study Phase III

107

Ari Seppänen Kai Steffens

Ministry of Environment PROBIOTEC GmbH

P.O. Box 399 Schillingsstrabe 333

00121 Helsinki D 52355 Düren-Gürzenich

Finland Germany

tel: +358/9-199-197-15 tel: 49/2421-69090

fax: +358/9-199-196-30 fax: 49/2421-690961

Robert Siegrist

Colorado School of Mines Rainer Stegmann

Environmental Science and Engineering Division Technische Universität Hamburg-Harburg

1500 Illinois Avenue Harburger Schlobstrabe 37

Golden, Colorado 80401-1887 D-21079 Hamburg

United States Germany

tel: 303-273-3490 tel: 49/40-7718-3254

fax: 303-273-3413 fax: 49/40-7718-2375

email: rsiegris@mines.edu e-mail: stegmann@tuharburg.d400de

Inge-Marie Skovgård Jan Švoma

Contaminated Land Division Aquatest a.s.

Danish Environmental Protection Agency Geologicka 4

29 Strandgade 152 00 Prague 5

DK-1401 Copenhagen K Czech Republic

Denmark tel: 420/2-581-83-80

tel: 45/3-266-0100 - direct 45/32660397 fax: 420/2-581-77-58

fax: 45/3-296-1656 e-mail: aquatest@aquatest.cz

e-mail: ims@mst.dk

Michael Smith Eberhard-Karls Universität - Tübingen

68 Bridgewater Road Geologisches Institut

Berkhamsted, Herts, HP4 1JB Sigwartstr. 10

United Kingdom 72076 Tübingen

tel: 44/1442-871-500 Germany

fax: 44/1442-870-152 tel: 49/7071-29-76468

e-mail: michael.a.smith@btinternet.com fax: 49/7071-5059

Marek Stanzel Kahraman Ünlü

KAP s.r.o. Depratment of Environmental Engineering

Skokanská 80 Middle East Technical University

169 00 Prague 6 Inönü Bulvari

Czech Republic 06531 Ankara

tel: 420/2-2431-3630 Turkey

fax: 420/2-5721-1255 tel: 90-312-210-1000

e-mail: kappraha@login.cz fax: 90-312-210-1260

e-mail: info@probiotec.ac-euregio.de

Gerhard Teutsch

e-mail: kunlu@rorqual.cc.metu.edu.tr
Treatment Walls NATO/CCMS Pilot Study Phase III

108

H. Johan van Veen Holger Weiss

The Netherlands Integrated Soil Research UF2 – Umweltforschungszentrum

Programme Leipzig-Halle GmbH

P.O. Box 37 Permoserstr. 15

NL-6700 AA Wageningen 04318 Leipzig

The Netherlands Germany

tel: 31/317-484-170 tel: 49/341-235-2060

fax: 31/317-485-051 fax: 49/341-235-2126

e-mail: anneke.v.d.heuvel@spbo.beng.wau.nl

Pál Varga Emergencies Engineering Division

National Authority for the Environment Environment Canada

Fö u.44 3439 River Road

H-1011 Budapest Ottawa, Ontario, K1A OH3

Hungary Canada

tel: 36/1-457-3530 tel: 613/991-1841

fax: 36/1-201-4282 fax: 613/991-1673

e-mail: vargap@kik.ktm.hu e-mail: harry.whittaker@etc.ec.gc.ca

Timothy Vogel Wolfgang Wüst

ATE/Rhodia Eco Services Institut für Wasserbau, Lehrstuhl für Hydraulik

17 rue Périgord und Grundwasser

69330 Meyzieu University of Stuttgart

France Pfaffenwaldring 61

tel: 33/4-7245-0425 70550 Stuttgart

Fax: 33/4-7804-2430 Germany

e-mail: timothy.vogel@rhone-poulenc.com tel: 49/711-685-4714

Harry Whittaker

fax: 49/711-685-7020

e-mail: ww@iws.uni-stuttgart.de