Permeable Reactive Barrier

Technologies for

Contaminant Remediation

United States

Environmental

Protection Agency

Office of Research and

Development

Washington DC 20460

EPA/600/R-98/125

September 1998

Office of Solid Waste and

Emergency Response

Washington DC 20460

Waste

Permeable Reactive Barrier

Water Table

Plume Treated Water

GW Flow
i

PERMEABLE REACTIVE BARRIER

TECHNOLOGIES FOR CONTAMINANT REMEDIATION

Robert M. Powell

Powell & Associates Science Services

Las Vegas, Nevada

David W. Blowes

University of Waterloo

Waterloo, Ontario

Robert W. Gillham

University of Waterloo

Waterloo, Ontario

Dale Schultz

Dupont Company

Newark, Delaware

Timothy Sivavec

GE Corporate Research and Development Center

Schenectady, New York

Robert W. Puls

U.S. EPA National Risk Management Research

Laboratory

Ada, Oklahoma

John L. Vogan

EnviroMetal Technologies Inc.

Guelph, Ontario

Patricia D. Powell

Powell & Associates Science Services

Las Vegas, Nevada

Rich Landis

Dupont Company

Wilmington, Delaware

Work Assignment Manager

Robert W. Puls

Subsurface Protection and Remediation Division

National Risk Management Research Laboratory

Ada, Oklahoma 74820

Technical Innovation Office

Office of SolidWaste and Emergency Response

U. S. Environmental Protection Agency

Washington DC 20460

National Risk Management Research Laboratory

Office of Research and Development

U. S. Environmental Protection Agency

Cincinnati, Ohio 45268

EPA/600/R-98/125

September 1998
ii

Notice

The U.S. Environmental Protection Agency through its Office of Research and Development funded and

collaborated in the research described here as indicated through the authors’ affiliations. ManTech Environmental

Research Services Corporation is an in-house support contractor to NRMRL/SPRD under 68-C3-0322 (Roger L.

Cosby, Project Officer). Dynamac is an off-site support contractor to NRMRL/SPRD under 68-C4-0031(David S.

Burden, Project Officer). Powell and Associates prepared the document as a subcontractor to ManTech

Environmental Research Services Corporation. Dupont and General Electric provided additional support by

providing time for Mr. Rich Landis, Dr. Dale Schultz, and Dr. Tim Sivavec to work on this report. This report has

been subjected to the Agency’s peer and administrative review and has been approved for publication as an EPA

document. Mention of trade names or commercial products does not constitute endorsement or recommendation

for use.

The Remedial Technology Development Forum (RTDF) provided support by way of authorship and document

review. All authors are members of the Permeable Reactive Barriers Action Team of the RTDF. Members of the

steering committee of the Permeable Reactive Barriers Action Team of the RTDF, not involved in authorship,

reviewed this document as well as members of the U. S. EPA’s Ground Water and Engineering Forums. Specific

individuals who reviewed the document and provided extensive comments which significantly enhanced the

technical quality included Mr. John Vidumsky (Dupont), Mr. Don Marcus, Mr. Mat Turner (New Jersey

Department Environmental Protection), Mr. Steve White (U.S. Army Corps of Engineers), Major Ed Marchand

(U.S. Air Force, AFCEE), and Mr. Peter McMahon (USGS).

All research projects making conclusions or recommendations based on environmentally related measurements

and funded by the U.S. Environmental Protection Agency are required to participate in the Agency Quality

Assurance Program. This project did not involve environmental related measurements and did not require a Quality

Assurance Plan.
iii

FOREWORD

The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation’s land, air, and

water resources. Under a mandate of national environmental laws, the Agency strives to formulate and implement

actions leading to a compatible balance between human activities and the ability of natural systems to support and

nurture life. To meet these mandates, EPA’s research program is providing data and technical support for solving

environmental problems today and building a science knowledge base necessary to manage our ecological resources

wisely, understand how pollutants affect our health, and prevent or reduce environmental risks in the future.

The National Risk Management Research Laboratory (NRMRL) is the Agency’s center for investigation of technological and management approaches for reducing risks from threats to human health and the environment. The

focus of the Laboratory’s research program is on methods for the prevention and control of pollution to air, land,

water, and subsurface resources; protection of water quality in public water systems; remediation of contaminated

sites and ground water; and prevention and control of indoor air pollution. The goal of this research effort is to

catalyze development and implementation of innovative, cost-effective environmental technologies; develop scientific and engineering information needed by EPA to support regulatory and policy decisions; and provide technical

support and information transfer to ensure effective implementation of environmental regulations and strategies.

Environmental scientists are generally familiar with the concept of barriers for restricting the movement of contaminant plumes in ground water. Such barriers are typically constructed of highly impermeable emplacements of

materials such as grouts, slurries, or sheet pilings to form a subsurface “wall”. The goal of such installations is to

eliminate the possibility that a contaminant plume can move toward and endanger sensitive receptors such as

drinking water wells or discharge into surface waters. Permeable reactive barrier walls reverse this concept of

subsurface barriers. Rather than serving to constrain plume migration, permeable reactive barriers (PRB’s) are designed as preferential conduits for the contaminated ground water flow. A permeable reactive subsurface barrier is

an emplacement of reactive materials where a contaminant plume must move through it as it flows, typically under

natural gradient, and treated water exits on the other side. The purpose of this document is to provide the most recent

information on PRB technologies in a format that is useful to stakeholders such as implementors, state and federal

regulators, Native American tribes, consultors, contractors, and all other interested parties. It includes information

on treatable contaminants, design, feasibility studies, construction options, site characterization needs and monitoring, as well as summaries of several current installations. It is hoped that this will prove to be a very valuable

technical resource for all parties with interest in the implementation of this innovative, passive, remedial technology.

Clinton W. Hall, Director

Subsurface Protection and Remediation Division

National Risk Management Research Laboratory
iv

Table of Contents

1.0 Executive Summary ............................................................................................................................ 1

2.0 Introduction.......................................................................................................................................... 4

2.1 Defining Permeable Reactive Subsurface Barriers ............................................................................. 4

2.2 Research Support and Application of The PRB Technology ............................................................. 5

2.3 Regulatory Acceptance of PRB Technology ....................................................................................... 6

3.0 Treatable Contaminants, Reactants and Reaction Mechanisms ......................................................... 8

3.1 Desirable Characteristics of Reactive Media ...................................................................................... 8

3.2 Treatable Contaminants ....................................................................................................................... 8

3.2.1 Halogenated Organic Compounds and Iron .................................................................................. 8

3.2.2 Redox-sensitive Inorganic Anions and Iron................................................................................ 11

3.2.3 Biologically Mediated Reduction and Removal of Anions ........................................................ 13

3.2.4 Adsorption and Precipitation of Inorganic Anions ..................................................................... 14

3.2.5 Reduction of Inorganic Cations .................................................................................................. 14

3.2.6 Biologically Mediated Reduction and Precipitation of Cations ................................................. 15

3.3 Enhancing Iron Reaction Rates ......................................................................................................... 15

4.0 Remediation Feasibility, Laboratory Treatability, and PRB Design Studies ................................... 16

4.1 Laboratory Treatability Studies ......................................................................................................... 16

4.1.1 Batch Studies ............................................................................................................................... 16

4.1.2 Column Studies ........................................................................................................................... 18

4.1.2.1 Column Test Methodology ............................................................................................. 18

4.1.2.2 Interpretation of Column Data ....................................................................................... 18

4.1.2.3 Interpretation of Changes in Non-Contaminant Inorganic Constituents ....................... 20

4.2 Determination of Required Residence Time in PRBs....................................................................... 22

4.2.1 Chlorinated Volatile Organic Compounds (VOCs) .................................................................... 22

4.2.2 Inorganic Constituents ................................................................................................................ 22

4.3 Ancillary Laboratory Studies............................................................................................................. 23

5.0 Site Characterization for Permeable Reactive Barriers .................................................................... 24

5.1 Hydrogeologic Characterization ........................................................................................................ 24

5.2 Contaminant Characterization ........................................................................................................... 25

5.3 Geochemical Characterization ........................................................................................................... 25

5.4 Microbial Aspects .............................................................................................................................. 26

5.5 Implementing the Field Characterization .......................................................................................... 27

6.0 Permeable Reactive Barrier Design .................................................................................................. 28

6.1 The Continuous PRB and Funnel and Gate Designs ........................................................................ 28

6.2 Emplacement Methods and Comparisons ......................................................................................... 30

6.2.1 Conventional Excavation ............................................................................................................ 30

6.2.2 Trenching Machines .................................................................................................................... 31

6.2.3 Tremie Tube/Mandrel ................................................................................................................. 31

6.2.4 Deep Soil Mixing ........................................................................................................................ 31

6.2.5 High-Pressure Jetting .................................................................................................................. 32

6.2.6 Vertical Hydraulic Fracturing and Reactant Sand-Fracturing .................................................... 32

6.3 Emplacement Verification ................................................................................................................. 32

7.0 Monitoring Permeable Reactive Barriers .......................................................................................... 33

7.1 Planning the Monitoring Effort ......................................................................................................... 33

7.2 Compliance Monitoring ..................................................................................................................... 33

7.2.1 Objectives .................................................................................................................................... 33

7.2.2 Compliance Monitoring System Design ..................................................................................... 33

7.2.3 Compliance Sampling Methods .................................................................................................. 34

7.2.4 Compliance Sampling Frequency ............................................................................................... 35

7.2.5 Contaminant Breakthrough/Bypass and Formation of Undesirable Products ............................ 36

7.2.6 General Water Quality Parameters ............................................................................................. 36
v

7.3 Performance Monitoring .................................................................................................................... 37

7.3.1 Objectives .................................................................................................................................... 37

7.3.2 Performance Monitoring System Design .................................................................................... 37

7.3.3 Performance Sampling Methods ................................................................................................. 38

7.3.4 Performance Sampling Frequency .............................................................................................. 38

7.3.5 Contaminant Degradation/Transformation ................................................................................. 39

7.3.6 Geochemical Indicator Parameters ............................................................................................. 39

7.3.7 Coring for Precipitate Buildup, Microbial Effects ..................................................................... 39

7.3.8 Hydrologic Testing for Permeability Alteration ......................................................................... 39

8.0 Field Installations .............................................................................................................................. 40

8.1 Chlorinated Hydrocarbon and Chromium Removal in Field Scale Systems .................................... 40

8.1.1 Industrial Site, Sunnyvale, California (January 1995) ............................................................... 40

8.1.2 Industrial Site, Mountainview, California (September 1995) .................................................... 40

8.1.3 Industrial Site, Belfast, Northern Ireland (December 1995) ...................................................... 43

8.1.4 Industrial Site, Kansas (January 1996) ....................................................................................... 43

8.1.5 USCG Facility, Elizabeth City, North Carolina (June 1996) ..................................................... 43

8.1.6 FHA Facility, Lakewood, Colorado (October 1996) .................................................................. 46

8.2 Interpretation of VOC Monitoring Data from the Field-scale Systems ............................................ 47

8.3 Inorganic Constituent Removal in Field-scale Systems .................................................................... 47

8.3.1 Nickel Rim Mine Site, Sudbury, Ontario (August, 1995) .......................................................... 47

8.3.2 Langton, Ontario On Site Wastewater Treatment (July, 1993) .................................................. 49

8.4 Biological Effects on Field-scale PRBs ............................................................................................ 50

8.4.1 Microbial Effects on Iron PRBs.................................................................................................. 50

8.4.2 Microbial Effects in PRBs for Inorganic Constituents ............................................................... 50

8.5 Effects of Mineral Precipitation on Field-scale PRB Performance .................................................. 51

Appendix A. Reference Table .......................................................................................................................... A-53

Appendix B. Explanation of Relevant Physical/Chemical Phenomena ........................................................... B-74

Appendix C. Scoping Calculations .................................................................................................................. C-78

Appendix D. Acronyms .................................................................................................................................... D-83

Appendix E. Glossary ......................................................................................................................................... E-86

References ............................................................................................................................................................ 89
vi

List of Tables

Table 1. Full-Scale in situ Remediation Projects. ............................................................................................ 5

Table 2. States That Have Approved the Installation of Iron PRBs. ............................................................... 6

Table 3. Contaminants Treatable By Reactive Materials in PRBs. ................................................................. 9

Table 4. Contaminants Presently Not Treatable By Fe(0). .............................................................................. 9

Table 5. Contaminants With Unknown Treatability. ...................................................................................... 10

Table 6. Typical Treatability Study Results, Elizabeth City Example. .......................................................... 17

Table 7. Construction Details of Six PRB Installations. ................................................................................. 41

Table 8. VOC Degradation Rate Data. ............................................................................................................ 49

Table 9. Required Weight-Per-Area (W/A) of Granular Iron. Calculated with n = 0.33,

P0

/P = 1000, and F = 3.5). Information and Equations from Appendix C. .................................... C-79

Table 10. Summary Information on Full-Scale Permeable Reactive Barriers Using Fe(0). .......................... C-80

Table 11. Cost Elements of Funnel and Gate Systems: High Estimate

(Gate @ $1000/yd3

; Funnel @ $25/ft2) ........................................................................................... C-82

Table 12. Cost Elements of Funnel and Gate Systems: Low Estimate

(Gate @ $200/yd3

; Funnel @ $10/ft2) ............................................................................................. C-82

List of Figures

Figure 1. Example of plume being treated by a permeable reactive barrier wall. ............................................ 1

Figure 2. Schematic of the column apparatus used in the bench-scale tests. ................................................... 19

Figure 3. Conceptual residence tme calculation for VOC removal. ................................................................ 21

Figure 4. Correlation of TCE degradation rates with temperature. .................................................................. 21

Figure 5. Potential effects of ground water flow diversion on plume direction. ............................................. 24

Figure 6. Effects of plume center-of-mass movement downgradient, and higher

concentrations impacting the barrier. Incomplete remediation is now occurring. ........................... 25

Figure 7. Potential effects of precipitation buildup on the iron over time. ...................................................... 26

Figure 8a. Plume capture by a continuous PRB trenched system.

The plume moves unimpeded through the reactive gate. ................................................................. 28

Figure 8b. Plume capture by a funnel-and-gate system. Sheet piling

funnels direct the plume through the reactive gate. .......................................................................... 28

Figure 9. Continuous trencher used at Elizabeth City, North Carolina. .......................................................... 34

Figure 10. Bundled tubes with short screens for performance monitoring. ....................................................... 38

Figure 11. PRB configuration, Sunnyvale, California. ....................................................................................... 40

Figure 12. PRB configuration, Mountainview, California. ................................................................................ 42

Figure 13. PRB vessel configuration, Belfast, Northern Ireland. ....................................................................... 44

Figure 14. PRB configuration, Kansas. .............................................................................................................. 44

Figure 15. PRB configuration, USCG facility, Elizabeth City, North Carolina. ............................................... 45

Figure 16. PRB configuration, FHA facility, Denver, Colorado. ....................................................................... 46

Figure 17. Pilot scale PRB configuration, Colorado and New York. ................................................................ 48

Figure 18. Example of an electrochemical corrosion cell. ............................................................................... B-75
vii
1

1.0 Executive Summary

Perhaps no recent remedial technology has generated as much interest as the use of subsurface permeable reactive

barriers (PRBs). This is due to the perceived PRB cost/benefit ratio and the potential of PRBs to mitigate the spread

of contaminants that have proven difficult and expensive to manage with other cleanup methods. The concept of a

PRB is relatively simple. Reactive material is placed in the subsurface where a plume of contaminated ground water

must move through it as it flows, typically under its natural gradient (creating a passive treatment system) and

treated water comes out the other side (Figure 1). The PRB is not a barrier to the water, but it is a barrier to the

contaminant. When properly designed and implemented, PRBs are capable of remediating a number of contaminants to regulatory concentration goals. It is currently believed that these systems, once installed, will have

extremely low, if any, maintenance costs for at least five to ten years. There should be no operational costs other

than routine compliance and performance monitoring.

The majority of installed PRBs use iron metal, Fe(0), as the reactive media for converting contaminants to nontoxic or immobile species. Iron metal has the ability to reductively dehalogenate hydrocarbons, such as converting

trichloroethene (TCE) to ethene. It can also reductively precipitate anions and oxyanions, such as converting

soluble Cr(VI) oxides to insoluble Cr(III) hydroxides. Organic materials are being used as reactive media in some

PRBs to biologically remediate certain other contaminants, such as nitrate and sulfate. Both laboratory and field

results have shown that the rate of transformation of these and many other contaminants is sufficiently rapid for

PRBs to be successfully used as full-scale remediation systems. Numerous other reactive materials are being

investigated, as are means to enhance both the iron and biological reactions. Many of the references for these

investigations are listed and described in Appendix A.

Commercial PRBs are currently built in two basic configurations (although others are being evaluated), the funneland-gate and the continuous PRB. Both have required some degree of excavation and been limited to fairly shallow

depths of fifty to seventy feet or less. Newer techniques for emplacing reactive media, such as the injection of

slurries, hydrofracturing, driving mandrels, etc., may serve to overcome some of these emplacement limitations.

The funnel-and-gate design PRB uses impermeable walls (sheet pilings, slurry walls, etc.) as a “funnel” to direct the

contaminant plume to a “gate(s)” containing the reactive media, whereas the continuous PRB completely transects

the plume flow path with reactive media. Due to the funnels, the funnel-and-gate design has a greater impact on

altering the ground-water flow than does the continuous PRB. In both designs it is necessary to keep the reactive

zone permeability equal to or greater than the permeability of the aquifer to avoid diversion of the flowing waters

around the reactive zone.

Figure 1. Example of plume being treated by a permeable reactive barrier wall.

Waste

Permeable Reactive Barrier

Water Table

Plume Treated Water

GW Flow
2

Several important issues must be addressed when considering contaminant remediation through the use of PRB

technology. These include the nature of the contaminant and the availability of reactive media that can transform

the contaminant yet remain reactive, in situ, for relatively long time periods. For contaminants of unknown

treatability or media of unknown reactivity, addressing these issues will require laboratory studies using both batch

and column techniques. The mobility, toxicity and stability of the transformation products resulting from the

contaminant and media interactions must also be assessed. If these transformation products are regulated

compounds, they must not exit the reactive zone of the PRB without themselves being immobilized or transformed

to innocuous compounds.

A thorough understanding of system hydrogeology and plume boundaries is needed prior to implementing a PRB,

due to the need for the plume to passively flow through the reactive zone of the PRB. The hydrogeologic

characterization must also yield information suitable for determining the rate of ground-water flow through the

reactive zone of the PRB. This is necessary to establish the ground-water/contaminant residence time per unit

thickness of reactive media which, when combined with the contaminant transformation rate as it passes through

the media, determines the total thickness of reactive media that is required. During PRB installation the reactive

media must be made accessible to the contaminant by some emplacement method and, as with most remedial

technologies, this becomes increasingly difficult at greater contaminant depth or for contaminants in fractured rock.

Once installed, the PRB should be carefully monitored for both compliance and performance; compliance to

ascertain that regulatory contamination goals are being met, and performance to assess whether the PRB

emplacement is meeting its design criteria and longevity expectations.

As for any remedial technology, it is important to fully understand the factors that can result in either successful

implementation and remediation or failure to achieve the remedial design goals. This document addresses the

factors, such as those mentioned above, that have been found to be relevant for successfully implementing PRBs for

contaminant remediation. Additionally, it provides sufficient background in the science of PRB technology to allow

a basic understanding of the chemical reactions proposed for the contaminant transformations that have been

witnessed both in the laboratory and in field settings. It contains sections on PRB-treatable contaminants and the

treatment reaction mechanisms, feasibility studies for PRB implementation, site characterization for PRBs, PRB

design, PRB emplacement, monitoring for both compliance and performance, and summaries of several field

installations. The appendices supplement this information with a detailed table of information available in the

literature through 1997, summarizing the significant findings of PRB research and field studies (Appendix A), a

further examination of the physical and chemical processes important to PRBs, such as corrosion, adsorption, and

precipitation (Appendix B), and a set of scoping calculations that can be used to estimate the amount of reactive

media required and facilitate choosing among the possible means of emplacing the required amount of media

(Appendix C). Appendix D provides a list of acronyms and Appendix E a glossary of terms that are used within this

document.

The goal of this Issue Paper is to provide the most recent information available on PRB technologies and to do so

in a format that is useful to stakeholders such as implementors, state and federal regulators, Native American tribes,

consultants, contractors, and all other interested parties. Other documents are also available which address PRB

topics that are not discussed in detail in this report to avoid duplicative effort, such as regulatory issues related to

PRB technology and cost information. For example, the Interstate Technology and Regulatory Cooperation (ITRC)

Workgroup (Permeable Barrier Wall Subgroup) has prepared a document titled “Regulatory Guidance for

Permeable Barrier Walls Designed to Remediate Chlorinated Solvents” (ITRC, 1997) and the Environics Directorate, U.S. Air Force, has published “Design Guidance for Application of Permeable Barriers to Remediate Dissolved

Chlorinated Solvents” (Battelle, 1997). Documents on the costs of PRB technology are being prepared by U.S.

EPA’s Technology Innovation Office (TIO) and by its Office of Research and Development, National Risk

Management Research Laboratory (ORD-NRMRL). Several web sites also provide information about PRB

technology. Among these are:

RemedialTechnologies Development Forum

• http://www.rtdf.org

Ground-Water Remediation Technologies Analysis Center

• http://www.gwrtac.org
3

U.S. EPA’s Robert S. Kerr Environmental Research Center of the National Risk Management

Research Laboratory

• http://www.epa.gov/ada/eliz.html

EnviroMetal Technologies, Inc.

• http://www.beak.com/eti.html

Powell & Associates Science Services

• http://www.powellassociates.com

Oregon Graduate Institute of Science and Technology

• http://www.cgr.ese.ogi.edu/iron
4

2.0 Introduction

A great deal of money and effort has been spent on environmental restoration during the past 30 years. Significant

progress has been made improving air quality, dumps and landfills, and surface water quality, although challenges

still exist in these areas. Among the more difficult and expensive environmental problems, and often the primary

factor limiting closure of contaminated sites following surface restoration, is contaminated ground water. The most

common technology used for remediating ground water has been to pump the water and treat it at the surface.

Although still useful for certain remedial scenarios, the limitations of pump-and-treat technologies have recently

been recognized, along with the need for innovative solutions to ground-water contamination (Keely, 1989;

National Research Council, 1994).

One of the most promising of these innovative solutions is the use of permeable reactive barriers (PRBs) filled

with reactive material to intercept and decontaminate plumes in the subsurface. The concept of PRBs is relatively

simple. Reactive material is placed in the subsurface to intercept a plume of contaminated ground water which must

move through it as it flows, typically under its natural gradient, thereby creating a passive treatment system. As the

contaminant moves through the material, reactions occur that transform it to less harmful (nontoxic) or immobile

species.

Many reactive media combinations can be envisioned for use in PRBs and numerous media and mixtures of media

are being investigated for a variety of contaminants (Appendix A). As of this writing, iron metal, variously

designated as Fe0

, Fe(0), or zero-valent iron, is the most common reactive media in the majority of field-scale and

commercial implementations. Scrap iron is not expensive and can be obtained in a granular form in the large

quantities needed. It has the ability to reduce oxidized inorganic species and reductively dehalogenate hydrocarbon

compounds. Reactive iron barriers depend upon corrosion to drive these reactions (Appendix B). For example,

chromate plumes are reduced from Cr(VI) to Cr(III) and precipitated, in situ, as chromium (III) hydroxides or

chromium-iron hydroxide solid solutions. An example from among the amenable halocarbon plumes are those

resulting from the dense, nonaqueous phase liquid (DNAPL) halogenated hydrocarbons. These include chlorinated

ethenes such as perchloroethylene (PCE), trichlorethylene (TCE), dichloroethylene (DCE), and vinyl chloride

(VC). The reductive dehalogenation of these compounds occurs due to electron transfers from the iron to the

halocarbon at the iron surface. This results in the halogen ions (e.g., Cl-

) being replaced by hydrogen species,

ultimately yielding ethene or ethane that can be mineralized via biodegradation.

The ability to perform dehalogenation in a reactive barrier is significant since the sources of DNAPL contamination, such as residual saturation, often cannot be easily located and may continue to generate a continuous plume of

dissolved halocarbons (National Research Council, 1994). Although these plumes can often be controlled by pumpand-treat, the systems require extensive maintenance and continual energy input. PRBs are also designed for plume

control, but with significant differences from pump-and-treat systems. PRBs are in situ systems, are intended to

operate in a totally passive manner, do not routinely bring the contaminant to the surface, and should operate for

years with minimal, if any, maintenance.

Although simple in concept, there is a great breadth of science and technology involved in the selection of reactive

materials for different contaminants and in the design, installation, and monitoring of these emplacements in the

subsurface. The purpose of this document is to provide an introduction and a guide to the science and technology of

PRBs.

2.1 Defining Permeable Reactive Subsurface Barriers

Environmental scientists are generally familiar with the concept of barriers for restricting the movement of

contaminant plumes in soil and ground water. Such barriers are typically constructed of highly impermeable

emplacements of materials such as grouts, slurries, or sheet pilings to form a subsurface “wall.” The goal of such

barriers is to minimize the possibility that a contaminant plume can move toward and endanger sensitive receptors,

such as drinking water wells, or discharge into surface waters. Rather than serving to constrain plume migration,

PRBs are designed as conduits for the contaminated ground-water flow. As contaminated water passes through the

reactive zone of the PRB, the contaminants are either immobilized or chemically transformed to a more desirable

(e.g., less toxic, more readily biodegradable, etc.) state. Therefore, a PRB is a barrier to contaminants, but not to

ground-water flow. A permeable reactive subsurface barrier is defined as:

an emplacement of reactive materials in the subsurface designed to intercept a contaminant plume, provide

a flow path through the reactive media, and transform the contaminant(s) into environmentally acceptable

forms to attain remediation concentration goals downgradient of the barrier (Powell and Powell, 1998;

Powell and Puls, 1997a).
5

In addition to PRBs, research is being conducted on a similar class of subsurface remediation technologies which,

in this document, will be referred to as reaction zone formation (RZF) technologies. These are somewhat different

from PRBs because they do not necessarily emplace solid phase reactive media in the subsurface. One approach, for

example, is the injection and withdrawal of a sodium dithionite solution into an aquifer matrix (Fruchter et al.,

1996). The dithionite reduces Fe3+ that is naturally available on the mineral matrix surface to Fe2+, thus increasing

the reducing capacity of the aquifer material itself. This has been shown to be an effective means of removing

chromate from ground water by reductive precipitation as the Cr(III) hydroxide, in much the same manner as was

previously discussed for its precipitation by Fe(0). Another RZF technology is the use of in situ electrodes, not to

directly move or remediate the contaminants, but rather to supply hydrogen as an electron donor to the subsurface

microbes. The microbes can then potentially remediate the contaminant, e.g., a chlorinated hydrocarbon, by using

it as the electron acceptor. This would reductively dehalogenate the compound, as was described above for the

electron transfer between Fe(0) and PCE. Although these RZF technologies are not the subject of this document, it

is important to realize that such processes are being investigated and implemented in the field.

2.2 Research Support and Application of the PRB Technology

The U.S. Environmental Protection Agency has supported the development of this innovative in-situ technology

through active collaboration on research involving the National Risk Management Research Laboratory (NRMRL)

and the National Exposure Research Laboratory (NERL) of USEPA’s Office of Research and Development (ORD),

through the Remediation Technologies Development Forum (RTDF) Permeable Barriers Action Team, and from

support provided by USEPA’s Technology Innovation Office (TIO). In addition, support has been provided from

several regional offices where sites are testing the technology at pilot-scale. Both DOD and DOE have also actively

supported research on PRB technology, financially supporting both laboratory and field trials.

As with any emerging technology, the field installation of this technology is limited relative to the research and

laboratory testing that has been performed and the overall potential of the technology for contaminant remediation.

Most full-scale PRBs are currently being used for the treatment of plumes of chlorinated hydrocarbons and

chromate, using granular Fe as the reactant. Laboratory studies have shown, however, that many other contaminants

can also be treated using PRB technology with the selection of the proper reactant (Appendix A). As of this writing

(March 1998), at least thirteen full-scale reactive barriers have been installed in the field (Table 1).

Table 1. Full-Scale in situ Remediation Projects

Industrial facility Sunnyvale, California gate and slurry wall December 1994

Mine Site, Sudbury, Ontario continuous wall August 1995

Industrial facility Mountainview, California gate and HDPE liner December 1995

Industrial facility Belfast, Northern Ireland in situ reactive vessel and slurry

wall

December 1995

Industrial facility Coffeyville, Kansas gate and slurry wall January 1996

USCG facility Elizabeth City, North

Carolina

continuous reactive wall June 1996

Government facility Lakewood, Colorado sheet pile funnels and 4 gates October 1996

Industrial facility South Carolina continuous iron/sand reactive

wall

October 1997

Industrial facility Colorado gate and slurry funnel November 1997

Industrial facility Oregon gates (2) and slurry funnel November 1997

Industrial facility Upstate New York continuous wall December 1997

Superfund site New Jersey continuous wall -

hydrofracturing

April 1998

U.S. DOE facility Kansas City, Missouri continuous wall April 1998
6

2.3 Regulatory Acceptance of PRB Technology

The USEPA recognizes this technology as having the potential to effectively remediate subsurface contamination

at many types of sites with significant cost savings compared to more traditional approaches (e.g., pump-and-treat).

The USEPA is actively involved in the evaluation and monitoring of this new technology to answer questions and

provide guidance to various stakeholder groups.

The Interstate Technology and Regulatory Cooperation (ITRC) Workgroup (Permeable Barrier Wall Subgroup) is

also actively involved in defining the regulatory implications associated with the installation of permeable reactive

barriers in the subsurface and in providing guidance on regulatory issues where possible.

The first full-scale commercial PRB was approved for use in the State of California by the San Francisco Regional

Water Quality Control Board (RWQCB) in 1994. Since that time, ten other full-scale systems have been installed

in the U.S., along with one in Sudbury, Canada, and one in Belfast, Northern Ireland. Table 2 lists some of the states

where permeable barriers have been approved for installation, either at pilot-scale or full-scale. The regulatory

approach to the technology at several sites has been to treat the installations as “at risk” remedies. That is, the owner

would be required to implement a more conventional remedy if the permeable barrier failed to meet performance

criteria. In some cases, the lack of existing groundwater use, particularly in industrial areas, facilitated implementation of PRBs. At sites in California and Colorado, there was a consensus among the regulators, site owners and

site consultants that existing pump and treat systems were not achieving the desired level of groundwater

remediation, making PRBs a favorable alternative solution.

From a federal perspective, one of the more significant advances for PRB technology occurred when a “chemical

treatment wall” was identified in June 1995 as the preferred alternative in the Record of Decision (ROD) at a

Superfund site (the Somersworth Municipal Landfill in Somersworth, New Hampshire). This decision, coupled

with the 1996 directive from the U.S. EPA Office of Solid Waste and Emergency Response (OSWER) to evaluate

alternatives to pump and treat remedies has encouraged consideration of the technology at other Superfund sites.

Active research programs at several U.S. EPA laboratories have also led to greater acceptance of the technology at

the federal level. A pilot-scale PRB for removal of VOCs in groundwater was evaluated in 1995 under the

Superfund Innovative Technologies Evaluation (SITE) program. The Technology Evaluation Report for this project

Table 2. Some States that have Approved the Installation of Iron PRBs.

State Full-Scale Installations Pilot-Scale Installations

California 2 2

Colorado 2 1

Delaware 0 1

Florida 0 3

Kansas 1 0

Illinois 0 1

Missouri 1 0

New Hampshire 0 1

New Jersey 1 0

New York 1 1

North Carolina 1 1

Oregon 1 0

South Carolina 1 1

Totals 11 11
7

remains to be published at the time of this writing (March 1998). The inclusion of PRBs in the SITE program and

Vendor Information System for Innovative Treatment Technologies (VISITT) has also increased the awareness of

the technology. Recent developments in interstate environmental technology verification programs may further

expedite regulatory acceptance.

In summary, regulatory acceptance of PRBs is expected to increase as the number of site installations increases

and more long-term performance data becomes available from existing installations.
8

3.0 Treatable Contaminants, Reactants and Reaction Mechanisms

3.1 Desirable Characteristics of Reactive Media

Reactive media used in permeable barriers should be compatible with the subsurface environment. That is, the

media should cause no adverse chemical reactions or byproducts when reacting with constituents in the contaminant

plume, and should not act as a possible source of contaminants itself. This requires that the material be well

understood and characterized. To keep PRB costs to a minimum, the material should persist over long periods of

time, i.e., it should not be readily soluble or depleted in reactivity, and the material should be readily available at a

low to moderate cost. The material selected should minimize constraints on ground-water flow by not having

excessively small particle size and it should not consist of a wide range of particle sizes that might result in blocked

intergranular spaces (i.e., it should preferably be unimodal with respect to grain size). Worker safety, with regard to

handling the material, should also be considered.

3.2 Treatable Contaminants

Table 3 lists contaminants that have been shown to be successfully treated by zero-valent iron or other media.

Tables 4 and 5 list some contaminants that have been shown to be not affected by zero-valent iron or have not yet

been fully evaluated, respectively.

3.2.1 Halogenated Organic Compounds and Iron

Considerable research during the past several years has focused on the degradation of chlorinated solvents, such as

TCE and PCE, by reactions at the surfaces of Fe(0). Although met with initial skepticism, the degradation process

is now widely accepted as abiotic reductive dehalogenation, involving corrosion of the Fe(0) by the chlorinated

hydrocarbon.

Iron corrosion processes in aqueous systems have been studied extensively. Until recently, the fate of corrosion

processes in dilute aqueous concentrations of chlorinated solvents acting as the oxidizing agents have not been

investigated. The net reductive dechlorination reaction promoted by Fe(0) (Equation 3) may be viewed as the sum

of anodic and cathodic reactions occurring at the iron metal surface (Equations 1 and 2, respectively), resulting in

hydrocarbon products if the dechlorination proceeds to completion.

Fe0 → Fe2+ + 2e- Anodic Reaction (1)

RCl + 2e-

 + H+ → RH + Cl- Cathodic Reaction (2)

Fe0

 + RCl + H+ → Fe2+ + RH + Cl- Net Reaction (3)

Under aerobic conditions, dissolved oxygen is usually the preferred electron acceptor and can compete with the

chlorinated hydrocarbon as the favored oxidant (Equation 4). Indeed, chlorinated hydrocarbons such as PCE and

carbon tetrachloride have oxidizing potentials very similar to that of O2

 (Archer and Harter, 1978). When sufficient

oxygen is present, the Fe2+ generated in Equation 4 further oxidizes to Fe3+ (Equation 5) and can precipitate as ferric

hydroxide or (oxy)hydroxides (Equation 6) at the elevated pH typical of corroding Fe systems. Corrosion of the iron

can generate large amounts of iron oxides and (oxy)hydroxide precipitates that can exert significant additional

chemical and physical effects within the reactive system (Powell et al., 1994; Powell et al., 1995a). The rapid

consumption of dissolved oxygen at the entrance to an iron system (column or barrier) has been shown to result in

these precipitates that might impact a system’s hydraulic performance at its upgradient interface (MacKenzie et al.,

1995; Mackenzie et al., 1997).

2Fe0

 + O2 + 2H2O → 2Fe2+ + 4OH- (4)

4Fe2+ + 4H+ + O2 → 4Fe3+ + 2H2

O (5)

Fe3+ + 3OH- → Fe(OH)3(s) (6)

Fe0

 + 2H2O → Fe2+ + H2 + 2OH- (7)

Fe2+ + 2OH- → Fe(OH)2(s) (8)

Anaerobic corrosion of iron by water (Equation 7) proceeds slowly. Both reactions 4 and 7 result in an increased

pH in weakly buffered systems, yielding ferric (oxy)hydroxides in aerobic systems (Equation 6) and ferrous

(oxy)hydroxides in anaerobic systems (Equation8). The aqueous corrosion of iron is mediated by the layer of

oxides, hydroxides and oxyhydroxides that are present at the iron-water interface. The formation of these

precipitates might further occlude the iron surface and affect its reduction-oxidation properties. However, this
9

Table 4. Contaminants Presently not Treatable by Fe(0)

Table 3. Contaminants Treatable By Reactive Materials in PRBs.

Organic

Compounds

Inorganic

Compounds

Methanes tetrachloromethane

trichloromethane

dichloromethane

Trace Metals Chromium

Nickel

Lead

Uranium

Technetium

Iron

Manganese

Selenium

Copper

Cobalt

Cadmium

Zinc

Ethanes hexachloroethane

1,1,1-trichloroethane

1,1,2-trichloroethane

1,1-dichloroethane

Anion

Contaminants

Sulphate

Nitrate

Phosphate

Arsenic

Ethenes tetrachloroethene

trichloroethene

cis-1,2-dichloroethene

trans-1,2-dichloroethene

1,1-dichloroethene

vinyl chloride

Propanes 1,2,3-trichloropropane

1,2-dichloropropane

Aromatics benzene

toluene

ethylbenzene

Other hexachlorobutadiene

1,2-dibromoethane

freon 113 

N-nitrosodimethylamine

Organic Compounds Inorganic Compounds

dichloromethane

1,2-dichloroethane

chloroethane

chloromethane

chloride

perchlorate
10

passive coating appears to be converted to magnetite (Odziemkowski et al. 1998), which is non-passivating, and

seems to allow sufficient contaminant degradation rates that can be sustained over years of operation in the ground.

Zero-valent iron is a mild reductant, and consequently the dehalogenation rates vary for the various chlorinated

solvents of environmental interest. A number of studies have shown that the primary determinant of degradation

rate is the specific surface area, or the surface area of iron per unit volume of pore water (Matheson and Tratnyek,

1994; Sivavec et al., 1995). Degradation rates are typically pseudo-first-order with respect to the halogenated

hydrocarbon, with the rate constant relatively insensitive to the initial hydrocarbon concentration. Studies of

published degradation rate data for individual halogenated hydrocarbons show that transformation rates are

proportional to iron surface area concentration (Johnson et al., 1996) and that observed rate constants can be

normalized to iron surface area to yield a specific rate constant, or ksa, for the halocarbon.

The reaction pathways by which Fe(0) reduces halogenated hydrocarbons have been determined for a few major

classes of chlorinated hydrocarbons. Such information is significant to the optimal design of a permeable reactive

barrier, as incomplete dechlorination of a highly chlorinated ethene, for example, could produce an intermediate

product, such as VC, which is more hazardous and more persistent than the parent compounds. Even very low

concentrations of undesirable by-products in the reactive barrier effluent must be avoided.

Typically, permeable reactive barriers are designed to provide adequate residence time in the treatment zone for

the degradation of the parent compound and all toxic intermediate products that are generated. At sites where the

ground-water contamination includes a mixture of chlorinated hydrocarbons, the design of the PRB is usually

determined by the least reactive constituent. Perhalogenated hydrocarbons tend to be reduced more rapidly than

their less halogenated congeners, and dechlorination is more rapid at saturated carbon centers (for example, carbon

tetrachloride) than at unsaturated carbons (for example, TCE or VC).

Excellent product mass balances have been determined for the transformation of several chlorinated ethenes (Orth

and Gillham, 1996) and methanes. These mass balance studies are best accomplished in a flow-through column

design, as sorption effects can be minimized. TCE, a common ground-water contaminant, is largely converted to

ethene and ethane. Generally less than 5%-10% of the initial TCE appears as chlorinated degradation products,

including the three dichloroethene products (cDCE is dominant) and VC. As shown in Equation 9, zero-valent iron

reduces TCE via two interconnected degradation pathways: (A) sequential hydrogenolysis and, (B) reductive ßelimination (Roberts et al., 1996). The intermediate products, cDCE and VC, are produced in the sequential

hydrogenolysis pathway, and are slower to degrade than is TCE itself. The chloroacetylene intermediate produced

via the ß-elimination pathway, by contrast, is a very short-lived intermediate and is very rapidly reduced to ethene.

The ß-elimination pathway accounts for the rapid conversion of TCE to ethene and ethane, with relatively minor

intermediate product formation.

Table 5. Contaminants with Unknown Treatability

Organic Compounds Inorganic Compounds

chlorobenzenes

chlorophenols

certain pesticides

PCBs

mercury
11

 (9)

The degradation pathway for chlorinated methanes is a much simpler one than that for chlorinated ethenes, as the

ß-elimination route is not available to the methane family. Carbon tetrachloride has been shown to undergo

sequential hydrogenolysis to chloroform and dichloromethane (Equation 10) (Matheson and Tratnyek, 1994). The

formation of chloroform accounted for about 70% of the carbon tetrachloride lost. Methylene chloride appeared,

after carbon tetrachloride decreased to the detection limit, and typically accounted for about 50% of the chloroform

lost. No further reaction of dichloromethane has been detected in an unamended zero-valent iron system.

(10)

3.2.2 Redox-sensitive Inorganic Anions and Iron

Negatively charged anions and oxyanions, the dissociation products of Lewis acids formed by the hydrolysis of

cations, are important ground-water contaminants. Elements which occur as anions or oxyanions under natural

ground-water conditions include arsenic, selenium, chromium, technetium and antimony. In addition, the dissolved

nutrient species nitrate and phosphate occur as anions, as does sulfate. Because of the negative charge, anionic

species are not attracted to negatively charged mineral surfaces, which are usually predominant in aquifers under

neutral pH conditions. This limited tendency for adsorption, and the high solubility of minerals containing

oxyanions, result in the potential persistence of high concentrations of these elements in aquifers. Treatability

studies, pilot-scale field trials and full-scale demonstration projects have been conducted for a series of anionic

inorganic contaminants, including SO4

2

, NO3

-

, and those containing Cr(VI), Se(VI), As(III), As(V), Tc(VII).

Treatment of chromate, CrO4

2- which contains Cr(VI), has been the most extensively tested and demonstrated of

the anionic inorganic contaminants. Chromium commonly occurs in two oxidation states, Cr(III) and Cr(VI), in the

environment. Trivalent Cr(III) is relatively non-toxic and a micronutrient. It forms sparingly soluble hydroxide

precipitates under conditions prevalent in most surficial aquifers and is also readily adsorbed by some minerals.

Cr(VI) is a known carcinogen, which forms relatively soluble precipitates, resulting in the persistence of relatively
12

high concentrations of dissolved Cr(VI) in affected aquifers (Palmer and Puls, 1994; Puls et al., 1995). Cr (VI) is

usually speciated as chromate, CrO4

2-, under typical ground-water pH and Eh conditions. This results in a high

degree of mobility for the Cr(VI) because anions are not readily adsorbed to the predominantly negatively-charged

aquifer materials. Treatments to remove Cr(VI) from ground water typically use reduction to Cr(III) and

precipitation of insoluble Cr(III) hydroxide precipitates.

A range of solid-phases containing reduced iron have been observed to promote the reduction and precipitation of

Cr(VI), including elemental iron (Gould, 1982; Bowers et al., 1986; Bostick et al., 1990; Blowes and Ptacek, 1992;

Powell et al., 1994; Powell et al., 1995a), iron-bearing oxyhydroxides (Eary and Rai, 1989) and iron-bearing

aluminosilicate minerals (Eary and Rai, 1989; Kent et al., 1994). Iron bearing reductants have been evaluated as

potential candidate materials to promote Cr(VI) reduction in reactive barriers. Comparison of reaction rates

observed using Fe(0), pyrite (FeS2

), and siderite (FeCO3), indicate that more rapid Cr(VI) reduction is attained

using Fe(0) (Blowes and Ptacek, 1992; Blowes et al., 1997). It has also been shown that the rates of Cr(VI)

reduction are dependent on the Fe(0) itself (i.e., how it was refined and manufactured, its level of impurities, etc.)

and whether certain aluminosilicate-containing aquifer materials are present and mixed with the iron (Powell et al.,

1995a; Powell et al., 1995b; Powell and Puls, 1997).

The overall reactions for the reduction of Cr(VI) by Fe(0) and the subsequent precipitation of Cr(III) and Fe(III)

oxyhydroxides are:

CrO4

2- + Fe0

 + 8H+ → Fe3+ + Cr3+ + 4H2O (11)

(1-x)Fe3+ + (x)Cr3+ + 2H2

O → Fe(1-x)CrxOOH(s) + 3H+ (12)

The extent and rate of Cr(VI) removal by elemental iron has been evaluated in laboratory batch tests and stirred

batch reactors (Blowes and Ptacek, 1992; Powell et al., 1995a), column tests (Blowes and Ptacek, 1992; Powell et

al., 1994; Blowes et al., 1997a), pilot-scale field trials (Puls et al., 1995) and a full-scale field demonstration

(Blowes et al., 1997b). The results of these experiments indicate that the rate of Cr(VI) reduction and precipitation

is sufficient for use in ground-water remediation systems. The removal of Cr(VI) from solution is accompanied by

a sharp decrease in the Eh, from initially oxidized conditions (Eh > 100 mV) to very reduced conditions

(Eh < -300 mV), and a sharp increase in the pH from initially near neutral conditions (6.5 < pH < 8.5) to more basic

conditions (pH > 9.5). Details of these reactions and the mechanisms of CrO4

2- reduction by Fe(0) have been

described (Powell et al., 1995a; Powell and Puls, 1997a).

Scanning probe microscopy of the reaction precipitates has indicated a one-to-one correspondence of the locations

of Cr(III) and Fe(III) hydroxide phases on the surface of reacted iron filings, indicating coprecipitation and the

likelihood of the formation of a solid solution phase of the general formula (Crx

, Fe1-x)(OH)3 (Powell et al., 1994;

Powell et al., 1995a). More detailed mineralogical studies indicate that the dominant reaction product derived from

the laboratory experiments is a mixed Fe-Cr oxyhydroxide phase with the mineral structure of goethite (FeOOH) as

shown in Equation 12. The distribution of Cr throughout the structure of this phase is variable, suggesting

incorporation through solid-solution substitution (Blowes et al., 1997b). X-ray Photoelectron Spectroscopy (XPS)

indicates that the Cr within the precipitate is exclusively in the Cr(III) oxidation state, and that Fe present in the

precipitate is in the Fe(III) oxidation state (Pratt et al., 1997). Auger Electron Spectroscopy indicates that the Fe(III)

and Cr(III) of the precipitate surface occur in the hematite (Fe2

O3) structure, that is distinct from the goethite

structure of the bulk phase, suggesting Cr(III) exolution and formation of a chromite-like (Cr2

O3) phase at the

precipitate surface (Pratt et al., 1997).

Reduction and precipitation of other anion-forming elements by elemental iron have also been investigated at the

laboratory scale. Laboratory experiments conducted to evaluate the treatability of Tc indicate rapid reduction using

elemental iron (Bostick et al., 1990; Del Cul et al., 1993; Clausen et al., 1995). Reduction of Se(VI) by elemental

iron, and precipitation of elemental selenium (Se0

) is favored thermodynamically. McRae et al. (1997) observed

removal of 1000 μg/L of Se(VI) from solution in laboratory batch experiments using elemental iron, and proposed

the reaction:

HSeO4

-

 + 3Fe0 + 7H+ → 3Fe2+ + Se0

(s) + 4H2

O (13)

Although removal of Se(VI) was observed, the secondary reaction product was not isolated or identified.

Arsenic commonly occurs as a dissolved species in two oxidation states, As(V) and As(III), and less commonly in

other oxidation states, including As(0), As(-I) and As(-II). The As(V) oxidation state occurs as H3

AsO4, and its

dissociation products. The As(III) oxidation state occurs as H3

AsO3 and its dissociation products. Although As(III),

which has greater health effects, is considered to be potentially more mobile in ground-water systems, As(V) is also
13

observed to be mobile in natural systems. In addition, the rate of arsenic transformation between the As(III) and

As(V) oxidation states occurs rapidly in response to changing environmental conditions. Arsenic reduction to the

As(0) oxidation state and precipitation of native arsenic, or incorporation into a secondary arsenic sulfide, has been

proposed as a potential treatment technique (McRae et al., 1997). In batch tests conducted using elemental iron,

McRae et al. (1997) observed rapid removal of As(V) from concentrations of 1000 μg/L to < 3 μg/L over a

two-hour period. Similar experiments using As(III) and mixtures of As(III) and As(V) indicated equally rapid

removal rates. The mechanisms resulting in arsenic attenuation, and the potential duration of treatment are being

evaluated by continuing experiments (McRae et al., 1997). The applicability of Fe(0) in PRBs for permanent As

removal and retention requires further investigation.

Reduction of NO3

-

 by Fe(0) has been observed to proceed rapidly, resulting in production of NO2

-

, and

subsequently ammonium ion (Cheng et al., 1997; Rahman and Agrawal, 1997; Till et al., 1998). The proposed

pathway for the overall reaction is:

4Fe(0) + NO3- + 10H+ → 4Fe2+ + NH4

+

 + 3H2O (14)

During laboratory studies, NO3

-

 reduction was observed to proceed through NO2

-

 to NH4

+ formation. The rate of

reduction was observed to be first order with respect to the Fe0

 substrate. In addition the rate was observed to

decline with increasing duration, particularly in the presence of chloride (Rahman and Agrawal, 1997). However,

NH4

+

 formation is not desirable since its MCL is lower than that of NO3

-

. Less NH4

+ is formed when Fe0

 is used in

conjunction with microbial consortia (Till et al., 1998). [See section on Biologically Mediated Reduction and

Removal]

3.2.3 Biologically Mediated Reduction and Removal of Anions

Biological processes affect the cycling of numerous elements, including nitrogen, sulfur, iron and manganese.

Treatment strategies employing these biologically mediated reactions have been proposed for direct treatment of

nitrate (Robertson and Cherry, 1995; Vogan, 1993; Till et al., 1998), sulfate (Blowes et al., 1994; Waybrant et al.,

1995), and for indirect removal of other anions through precipitation as sulfide phases.

A denitrification system for removal of nitrate from ground water affected by discharge from site wastewater

disposal systems has been developed (Robertson and Cherry, 1995). This system intercepts a plume of nitratebearing ground water with a reactive barrier containing solid phase organic carbon. In the presence of organic

carbon, under anaerobic conditions maintained below a water cover in the subsurface, reduction of NO3

-

 to N2 gas

is thermodynamically favored. The proposed reaction is:

5CH2

O(s) + 4NO3

- → 2N2

 + 5HCO3

-

 + 2H2O + H+ (15)

CH2

O represents a simple form of organic carbon which is catalyzed by bacteria of the Pseudomonas group. These

bacteria use NO3

-

 as an electron acceptor in the oxidation of organic carbon. In a laboratory study a series of

inexpensive organic carbon sources were evaluated to assess their potential use in reactive barriers for treatment of

NO3

- (Vogan, 1993). Results to date indicate that readily available materials, including sawdust and wood waste, are

suitable materials for a reactive barrier system. Robertson and Cherry (1995) evaluated permeable reactive barriers

for treating NO3

-

 at several domestic and institutional septic systems. The results of these studies indicate that

sufficient denitrification occurs rapidly and reduces NO3

-

 from concentrations typically observed in the effluent of

onsite wastewater disposal systems (5 - 90 mg/L NO3

- - N) to below the World Health Organization drinking water

standard (10 mg/L NO3

-

 - N). Reactive barriers using a variety of design configurations have been implemented. No

evidence of assimilatory NO3

-

 reduction and NH4

+

 formation has been observed.

A combined system of Fe(0) and the denitrifying bacteria Paracoccus denitrificans to reduce NO3

-

 has been

investigated (Till et al., 1998). The hydrogen produced by the corrosion of the iron was used as the electron donor

for nitrate-based respiration. Using steel wool as the Fe0

 they found that the combination increased nitrate removal

rates, relative to steel wool alone, and transformed a greater portion of the added NO3

-

 to innocuous gases rather

than to NH4

+

 .

Biologically mediated reduction of sulfate to sulfide, accompanied by the formation of metal sulfides occurs

through the reaction sequence

2CH2

O(s) + SO4

2- + 2H+ → H2

S(aq) + 2CO2(aq) + 2H2O (16)

Me2+ + H2

S(aq) → MeS(s) + 2H+ (17)

where CH2

O represents organic carbon and Me2+ represents a divalent metal cation in solution. Biologically

mediated sulfate reduction has been proposed to treat metal cations derived from mine sites in wetlands (McIntire
14

et al., 1990; Kleinmann et al., 1991; Eger and Wagner, 1995), bioreactors (Dvorak et al., 1992) and permeable

reactive barriers (Blowes et al., 1995; Waybrant et al., 1995; Benner et al., 1997; Blowes et al., 1997b). Although

these systems are designed to promote the removal of dissolved metals, metal removal is accompanied by removal

of sulfate. In laboratory studies, (Waybrant et al., 1995, 1997a) sulfate removal was observed at rates of 0.14 to

4.23 mg L-1 day- 1 g-1 of media.

3.2.4 Adsorption and Precipitation of Inorganic Anions

Inorganic anions which are not susceptible to reductive or oxidative processes must be removed from solution by

other means. These anions, as well as redox-sensitive species, may be removed by precipitation, adsorption or

coprecipitation on mineral surfaces.

Artiole and Fuller (1979) proposed the use of barriers containing crushed agricultural limestone to neutralize the

pH of landfill leachate and precipitate Cr. Laboratory tests indicated successful removal of the cation form, Cr3+,

and less successful treatment of the anionic form of Cr(VI), CrO4

2-. Thomson et al. (1991) and Longmire et al.

(1991) described the potential use of sphagnum peat, crushed limestone, and hydrated lime, to remove U, As, Mo,

and Se in laboratory batch tests. Morrison and Spangler (1992, 1993) proposed the use of a series of industrial

byproducts as reactants to remove U, As and Mo in precipitation or sorption barriers. Phases evaluated included

bases, reductants, and sorbents, including hydrated lime, ferrous sulfate, hydrous ferric oxides, peat, and phosphatebearing phases to remove both anionic and cationic species from solution. The system described by Morrison and

Spangler (1992; 1993) is potentially suitable for a wide range of anionic contaminants.

Phosphate (PO4

3-) derived from anthropogenic sources is a limiting nutrient in many aquatic ecosystems. Release

of excess phosphate results in accelerated biological activity, and ultimately in eutrophication of the aquatic

ecosystem. In many regions, phosphate released from onsite wastewater disposal systems, such as septic systems or

sewage lagoons, results in excess phosphorus. Baker et al. (1996, 1997) used a mixture of iron oxide, calcium oxide

and limestone to promote adsorption and coprecipitation of phosphorus. Phosphate was removed rapidly in batch

experiments. In column experiments extending over 3 years and 1200 pore volumes, influent PO4

3- , at a

concentration of 4 mg/L PO4

3- - P, was removed to <0.3 mg/L PO43- - P, indicating more than 90% removal of the

phosphate over a prolonged period. Field-scale demonstration at an institutional septic system resulted in removal

of phosphate from influent concentrations of 1 to 2 mg/L PO4

3- - P to less than 0.1 mg/L PO43- - P. Test systems,

established at onsite wastewater treatment systems, indicate potential for removal of PO4

3- from wastewater under

continuous flow conditions. The effluent concentrations achieved by this technique are sufficiently low to prevent

eutrophication of surface water flow systems in which PO4

3- is the limiting nutrient (Baker et al., 1996).

The mixture used by Baker et al. (1996) was evaluated for the potential treatment of arsenic (McRae, 1997). Batch

experiments indicated a decrease in As(III) from 1000 μg/L to < 3 μg/L in less than 2 hours. Experiments showed

decreases in As(V) concentrations, and in mixtures of As(III) and As(V) from 1000 μg/L to <3 μg/L over a similar

period. A column test conducted using this material indicated decreases in mixed As(III) and As(V) concentrations

from a total of 1000 μg/L to <3 μg/L for more than 120 pore volumes (McRae et al., 1997). Research on this system

is continuing.

3.2.5 Reduction of Inorganic Cations

Positively charged inorganic cations, including the metals Cd, Co, Cu, Mn, Ni, Pb, Zn, and complex cations such

as UO2

2+ are important ground-water contaminants. High concentrations of these metals are associated with

industrial wastes, mine wastes and nuclear waste disposal sites. Ground water within mine waste piles, and leachate

derived from mine wastes commonly contains high concentrations of dissolved metals (Dubrovsky et al., 1984;

Morin et al., 1988). Leachate derived from waste disposal areas, containing high concentrations of dissolved

uranium and technetium, has been reported from the Oak Ridge National Laboratory (Olsen et al., 1986).

Reduction and precipitation of sparingly soluble solids has been proposed for inorganic cations, as well as anions.

Treatment of cation-forming electroactive metals, including mercury, uranium, copper and technetium by reduction

with elemental iron and coprecipitation within secondary precipitates has been investigated at the laboratory scale.

Laboratory batch experiments indicate rapid removal of U(VI) and Tc(VII) from solutions in contact with elemental

iron (Bostick et al., 1990; Liang et al., 1996). The reaction between U(VI) and Fe(0) can be expressed as

Fe0

 + UO2

2+

(aq) → Fe2+ + UO2(s) (18)

where UO2(s) is an amorphous or crystalline uranium oxide precipitate. Strongly reducing conditions must be

attained for uranium reduction and precipitation to proceed. In addition, if oxidizing conditions recur, the reduced

uranium may become reoxidized and remobilized. Other metal cations potentially treatable by reduction with Fe(0)

include Cu and Hg.
15

3.2.6 Biologically Mediated Reduction and Precipitation of Cations

Biologically mediated reduction reactions can promote the attenuation of inorganic cations. These reactions

include direct reduction of the cation and precipitation of a sparingly soluble solid, and indirect precipitation

resulting from the oxidation or reduction of an inorganic anion. An example of direct reduction is the reduction of

U(VI) through a bacterially mediated reaction. Bacteria have been isolated that are capable of reducing U(VI) to

U(IV), and sustaining metabolism based on this reaction (Lovely et al., 1991; Lovely and Phillips, 1992a; Lovely

and Phillips, 1992b). It was found that U(VI) reduction by organic matter or H2

 was insignificant without microbes

present. The proposed reaction for catalysis of U(VI) reduction by acetate, using the microbe G. matallireducens is:

CH3

COO- + 4UO2(CO3)2

2- + 4H2

O → 4UO2 + 10HCO3

- + H+ (19)

It was also proposed that the A. putrefaciens could oxidize hydrogen to reduce U(VI) by the reaction:

H2

 + UO2(CO3)2

2- → UO2

 + 2HCO3

- (20)

Indirect precipitation of inorganic cations results from the reduction of an anion-forming species, usually sulfate.

Sulfate reduction generates hydrogen sulfide, which combines with metals to form relatively insoluble metal sulfide

precipitates, described in Equations 16 and 17. Laboratory studies indicate that many metals are treatable using this

approach, including Ag, Cd, Co, Cu, Fe, Ni, Pb, and Zn (Waybrant et al., 1995; Waybrant et al., 1997a). Column

experiments, conducted using a range of organic substrates demonstrated the potential to remove a range of

dissolved metals at ground-water velocities similar to those observed at sites of ground-water contamination. A

field-scale reactive barrier for the treatment of acid mine drainage and removal of dissolved Ni was installed in

1995, at the Nickel Rim mine site near Sudbury, Ontario. It was composed of municipal compost, leaf compost and

wood chips. Monitoring of the reactive barrier indicates continued removal of the acid generating capacity of the

ground water flowing through the PRB and decreases in dissolved Ni concentrations from up to 10 mg/L to

<0.1 mg/L within the PRB. Monitoring is continuing at this site.

3.3 Enhancing Iron Reaction Rates

The range of reactivity of halogenated hydrocarbons and other contaminants with zero-valent iron illustrates the

value of achieving faster degradation rates and more complete degradation of less reactive products. Research into

increasing these rates has investigated metals that are stronger reducing agents than iron, the addition of

aluminosilicate minerals, and metal couples. Metals that are stronger reducing agents than iron have higher

oxidation potentials; i.e., yield their electrons more readily to an oxidized substance, hence they corrode faster.

Aluminosilicate mineral addition and metal couples result in increased corrosion rates for the iron that is present.

Many aluminosilicate minerals seem to enhance corrosion of Fe(0) by buffering reaction pH to lower values

(pH=7-8 rather than > 9). This buffering comes from dissolution of the aluminosilicate minerals (Powell et al.,

1995a; Powell and Puls, 1997b) with the generation of hydrogen as protons. These protons can coordinate with the

cathodic regions of the iron surfaces, accepting electrons and thereby increasing corrosion rates. In laboratory batch

tests using Fe(0) with an aluminosilicate-containing aquifer material from Elizabeth City, North Carolina, the halflife of Cr(VI) in solution was decreased by an order of magnitude relative to a system with Fe(0) and silica sand.

The Cr(VI) half-life was about two orders of magnitude shorter than tests where only Fe(0) was present (Powell and

Puls, 1997). The protons (H+

) or their reduced species, surface coordinated monoatomic hydrogen atoms (H•), may

also serve as replacement moieties for the halogen atoms on the hydrocarbons that are being reductively

dehalogenated at the cathodic surfaces (Powell and Puls, 1997b).

Bimetallic systems (metal couples) prepared by plating a second metal onto a zero-valent iron surface, including

Fe/Cu, Fe/Ni (Sivavec et al., 1997; Gillham et al., 1997) and Fe/Pd (Muftikian et al., 1995), have been shown to

accelerate solvent degradation rates relative to untreated iron metal. Palladized iron has been shown to be effective

in dechlorinating halogenated aromatic compounds such as polychlorinated biphenyls (PCBs) in addition to

chlorinated aliphatic compounds (Grittini et al., 1995). The rate enhancement observed in bimetallic systems may

be attributed to corrosion-inducing effects promoted by the second, higher reduction potential metal and possibly

some catalytic effects. However, some investigators have found the enhanced reactivity of these systems to

diminish relatively quickly, whereas others have found no apparent loss of reactivity (Gillham et al, 1997). These

differences may be related to ground-water chemistry or the method used for plating the iron, but further

investigation is needed. It is important to note that zero-valent iron systems have not shown similar losses in

reactivity in long-term laboratory, pilot and field investigations.
16

4.0 Remediation Feasibility, Laboratory Treatability, and PRB Design Studies

Most, if not all, PRB installations have been designed and implemented based on the results of laboratory batch

and column studies used to test reactant materials and the kinetics of contaminant removal. These data are used in

combination with site-specific information (such as ground-water velocity, contaminant type and concentration,

and the total mass flux of the contaminant requiring treatment). The following sections present information on these

laboratory studies and how the information is used in the development of a PRB as a remedial solution. However,

there will often be a need to develop some approximate design configurations and cost estimations prior to carrying

out extensive laboratory studies, to determine whether PRB technology is more suitable for a site than other options.

Appendix C provides some scoping calculations that have been proposed to allow such cost estimation. These can

also provide some insight into the PRB emplacement methods and designs that could provide iron in thicknesses

sufficient to meet the regulatory requirements of the site.

4.1 Laboratory Treatability Studies

The need for laboratory treatability studies for PRB design is primarily dependent on the contaminants present,

their concentration and the geochemical conditions at the site. For contaminants where an extensive database of

removal rates exists (TCE or chromate in granular iron, for example), these rate data can be used for design

purposes and the treatability tests can potentially be omitted. When there are mixtures of contaminants, the

geochemical conditions are significantly different from sites previously tested, or where reactive mixtures or

sequential zones of reactive materials are proposed, treatability tests can be highly instructive and are strongly

recommended. Laboratory treatability studies can be used to compare the reactivity and longevity of reactive

materials under uniform and controlled conditions, as well as to estimate half-life (t1/2) information useful for PRB

design. In addition, performing studies at differing experimental temperatures allows determination of the

temperature dependence of the reaction rates. This temperature dependence might be described by an Arrhenius

expression, allowing prediction of rate constants for various types of reactions (i.e., abiotic reductive dechlorination

or biologically mediated contaminant removal) over a range of temperature.

Laboratory treatability studies should be conducted using ground water from the subject plume. Although VOC

degradation rates in granular iron are not greatly influenced by the inorganic chemistry of ground water,

information on inorganic geochemical changes has proven to be very useful. Removal of inorganic contaminants

may be strongly influenced by the background geochemistry of the plume. Every effort should be made to maintain

the oxidation-reduction (redox) state of the ground water used in the studies. This requires both proper sampling

and storage in the field and as the water is transferred from the field sample bottles to the influent reservoir. The

possible effects of sample storage on redox state can be evaluated by comparing pre-batch and column laboratory

results to field pH, DO, and Eh measurements.

4.1.1 Batch Studies

Batch treatability studies are most suitable for screening candidate reactive barrier materials. Results obtained

using various types of materials give relative rates that can be useful in selecting the most appropriate material(s)

for subsequent testing and/or field application. Batch tests are usually faster, cheaper, and simpler to set up than

columns, and allow rapid comparison of varied parameters on the experimental results (Powell et al., 1995a).

Laboratory batch treatability tests should include blanks, which contain only site water, and reactive samples

containing the candidate reactive materials along with the site water. It is simple, for example, to determine which

of two types of reactive media is most effective for remediating a contaminant using batch tests. Three or more

replicates for each of the two media types can be set up in capped tubes or bottles using a constant aqueous volume,

contaminant concentration, and media mass (or surface area). After shaking for some interval, the concentration of

the contaminant can be analyzed in each of the replicates and a determination made whether a difference exists in

the effectiveness of the two types of media.

Increasing the experimental complexity somewhat by adding multiple sampling intervals allows the determination

of the rate of contaminant removal using the equations of classical kinetics. Batch systems are somewhat limited in

this regard, however, since shaking of the tubes or bottles negates many of the mass transport and diffusive effects

that would limit reactions in unshaken systems. Also, batch tests usually have very low ratios of reactive material

to solution relative to column tests and actual field implementations. Although Johnson et al. (1996) found that

batch and column studies did not give distinguishable distributions of results when rate data for VOCs was

normalized to iron surface area (kSA), they did observe that experimental conditions determined whether the

degradation rates were reaction or diffusion limited. To develop a table of representative kinetic data for

dehalogenation by iron metal, they averaged reaction rates for a given halocarbon from column studies, batch tests,
17

and regression analyses of observed rates (kobs) versus iron surface area concentration (i.e., m2 L-1). This resulted in

average kSA values with relatively large standard deviations for each halogenated organic compound. For example,

TCE had an average kSA value of (3.9±3.6) x 10-4. Therefore, some caution is warranted when results from

laboratory experiments are extrapolated to field-scale systems.

The test procedures can vary significantly depending upon contaminants of interest and characteristics of the

reactive material. For a site at the U.S. Coast Guard Air Support Center, Elizabeth City, North Carolina, a series of

batch experiments were conducted to evaluate the effectiveness of various commercial iron sources for the

simultaneous removal of both hexavalent chromium, Cr(VI), and trichloroethene (TCE) from the site water. The

results are shown in Table 6, together with the column test results using the same site water. From Table 6, it is

apparent that using TCE batch test half-lives would result in an overdesign of the field treatment system. However,

Table 6. Typical Treatability Study Results, Elizabeth City Example.

6.1 Trichloroethene Half-Lives from Batch Tests.

Iron Source (1, 2, or 3) t1/2 (hr)

1 plus silica sand 32

2 plus silica sand 73

3 plus silica sand 31

1 plus aquifer material 22

6.2 Time to Non-Detectable Chromium Levels in Batch Tests.

Iron Source (1, 2, or 3) time to non-detect (hr)

1 plus silica sand 0.6

2 plus silica sand 1.2

3 plus silica sand 0.4

1 plus aquifer material 0.72

6.3 Trichloroethene Half-Lives from Column Tests (v = + 2 ft/day).

Reactive Mixture t1/2 (hr)

100% Source 1 Iron 0.61

100% Source 3 Iron 0.80

50% Source 1 Iron, 50%

Silica sand

0.97

50% Source 1 Iron, 25%

sand, 25% aquifer material

0.48

50% Source 2 Iron, 25%

sand, 25% aquifer material

0.99

Note: chromium removal to detection limits was obtained within 0.1 ft of travel distance in all columns

(i.e., within ± 1.2 hrs)
18

the data does indicate which of the commercial mixtures tested (in this case iron sources 1 and 3) were the best

candidates for further testing using columns.

A second use of batch tests involves the evaluation of new PRB emplacement methods. Several existing

construction techniques, used historically to build low permeability (containment) walls, are being modified to

facilitate the construction of permeable iron barriers at greater depths. Several of these involve the use of a finer

grained fraction of iron in a biodegradable slurry mixture to allow injection or placement of this mixture in the

subsurface. Laboratory-scale and limited field-scale testing has shown that the reactivity of the iron is maintained

following contact of the material with guar based slurry; however, this should still be confirmed on a site-specific

basis. If an enzymatic or acid-based breaker is used to promote slurry breakdown, then batch tests can also be used

to confirm that the breaker used will function in the high pH, low Eh conditions established in the Fe(0).

4.1.2 Column Studies

Laboratory column tests are useful in determining contaminant removal rates under conditions that more closely

approximate the operating conditions anticipated in the field, such as flow velocity. These rates are the basis for the

design parameters used to determine the required residence time for the contaminant in the reactive material. Using

the residence time and the flow rate, the thickness of the treatment zone can be determined. The laboratory column

tests may also include sampling of the column profile and/or influent and effluent, in addition to the contaminant,

to assess changes in the major ion composition of the water. These data provide information concerning potential

mineral precipitation in the reactive material caused by changing redox potential (Eh) and pH conditions, which are

also important parameters affecting the removal of inorganic contaminants in PRBs. Mineral precipitation rates

may also impact the operation and maintenance requirements for VOC removal systems.

Though column tests are more costly and time consuming than batch tests, they typically yield more realistic field

performance rates, provide a better opportunity to examine products of the reactions and can provide useful

information concerning long-term performance.

4.1.2.1 Column Test Methodology

Columns are typically 10 to 100 cm long, and 2.5 to 3.8 cm inside diameter, with sampling ports at the influent and

effluent ends and preferably, also along the length (Figure 2). The sampling ports are designed to allow water

samples to be collected along the central axis. Details concerning column test methodology can be found in several

publications (see Appendix A).

Ground water obtained from the site is supplied to the influent end of the column at a constant flow velocity using

a laboratory pump. The flow velocity is selected to approximate the velocity expected in the field-scale treatment

zone. Contaminant concentrations are measured at the inlet, outlet, and sampling ports along the column every 5 to

10 pore volumes (one pore volume is equal to the total volume of liquid within the column) until a steady-state

concentration profile is achieved (i.e., concentrations at a point remain relatively constant over time). For example,

this generally occurs after 40 to 100 pore volumes of flow for VOCs in granular iron. Eh and pH profiles are also

measured periodically during the test period. Concentrations of the major cations, anions, and alkalinity are usually

monitored at less frequent intervals to help predict the potential for mineral precipitation within the reactive

material. If necessary, other chemical parameters relevant to a particular site can also be measured.

4.1.2.2 Interpretation of Column Data

For each test column at each velocity, contaminant concentrations are plotted as a function of distance along the

column. The flow rate is used to calculate the residence time at each sampling position (relative to the influent) for

each profile. The contaminant degradation or disappearance rate constants are calculated for each contaminant in

the influent solution ground water, using kinetic models. For VOCs and/or chromate a first-order model is used:

C = C0

e-kt (21)

where:

C = contaminant concentration in solution at time t,

C0

 = initial contaminant concentration of the influent solution,

k = first order rate constant, and

t = time.

By rearranging and taking the natural log, Equation (21) becomes:

ln (C/C0

) = -kt (22)
19

The time at which the initial concentration declines by one-half, (C/C0 = 0.5), is the half life, which, by rearranging

equation (22), is given by:

t

1/2 = 0.693/k (23)

The first-order rate constant, k, has the units of time-1 (e.g., sec-1, hr-1) and is computed from the slope of the firstorder model obtained by fitting equation (22) to the experimental data. Half lives, along with corresponding

standard error of estimate (sy|x

)values can then be calculated.

When the contaminant is a VOC, breakdown products from the VOCs in the influent solution (e.g., dichloroethene

isomers from TCE) may reach maximum concentrations at an interior sampling port. In this case, an approximate

degradation rate for the breakdown product can be calculated using the maximum concentration at this port, rather

than the influent concentration, as C0

. Ideally, both parent and intermediate concentration data should be fitted

using a first-order kinetic model to determine degradation rates and conversion factors.

A variety of mechanisms are potentially suitable for the removal of inorganic contaminants. Many of the

mechanisms employed in treatment of inorganic species rely on the precipitation or adsorption of a chemical

constituent. Laboratory batch and column data should be combined with geochemical modeling to assess the

stability of potential precipitates, adsorbates, and to assess the potential utility of reactive mixtures for remediation

of inorganics. Altering the concentrations of potential nutrients (e.g., NH4

+

, PO4

3-) in biologically mediated systems

permits an assessment of potential nutrient limitations on the rate of contaminant removal.

A comprehensive characterization of the water chemistry within a batch test permits the use of geochemical

speciation calculations. Geochemical models commonly used to perform these calculations include MINTEQA2

Solution

Reservoir

Pump

Influent

Sampling Port

Sampling

Ports

Plexiglass

Column

Effluent Samples

Figure 2. Schematic of the column appartus used in the bench-scale tests.
20

(USEPA, 1993) and PHREEQC (USGS, 1998). These models contain large and comprehensive databases that can

be used to evaluate the tendency for mineral phases to precipitate or dissolve based on the saturation state of

anticipated minerals. MINTEQA2 also includes a variety of surface ionization/complexation models, which can be

used to evaluate the potential for contaminant adsorption onto mineral surfaces.

The results from laboratory experiments can be augmented by the application of mineralogical characterization

techniques and surface analytical procedures. These techniques allow the isolation and identification of reaction

products, identification of the oxidation state of adsorbed or precipitated elements, and characterization of the

mineral structure of secondary precipitates. Mineralogical techniques can also be used to verify the identity of

secondary phases that are inferred from geochemical modeling. Accumulations of secondary minerals and the

occurrence of coatings can be used to infer reaction progress and to assess potential transport limitations arising

from reaction products blocking reactive surfaces.

When used in conjunction with geochemical speciation/mass transfer calculation, kinetics and column flow data

can also be used to assess the rate-determining steps within an overall reaction sequence. The results of long-term

column experiments can be used to assess variations in the reaction rates resulting from consumption of substrate

materials, growth of bacterial species, and the formation of potentially deleterious reaction products (Waybrant et

al., 1997a). These data can also be used to assess changes in the dominant reaction mechanisms.

The effectiveness of treatment of contaminant species may be limited by changes in the accessibility of the

reactive material due to the formation of coatings by secondary phases. The degree and duration of treatment may

also be limited by depletion of the mass of reactive material, or in the case of adsorbent barriers, by the availability

of reaction sites on the substrate surface. Continuing controlled column experiments for prolonged periods of time

provides an opportunity to drive the treatment system to failure, to evaluate the potential limitations of proposed

field installations (Blowes et al., 1997a).

4.1.2.3 Interpretation of Changes in Non-Contaminant Inorganic Constituents

The carbonate equilibrium present in the contaminant plume may be significantly affected as the ground water

passes through the reactive barrier. In Fe(0) columns, calcium and alkalinity concentrations normally decrease as

ground water passes through the iron in response to increasing pH caused by the corrosion of iron. Corrosion also

causes the Eh of the ground water to decrease (see Appendix B). As the pH of the solution increases, bicarbonate

(HCO3

-

) ions are converted to carbonate ions (CO3

2-). The CO32- ion formed can then combine with the cations

present in solution (Ca2+, Fe2+) to form carbonate mineral precipitates such as calcite (CaCO3

) and siderite (FeCO3).

At some sites, Mg2+ may precipitate in solid solution with CaCO3

.

Independent corrosion rate measurements of metallic iron (Reardon, 1995) indicate that several mmol/L Fe2+

would be introduced to ground water in these columns due to iron corrosion. In many instances, iron precipitates

including siderite (FeCO3

), iron hydroxides and oxyhydroxides, and “green rusts” will form in the column.

Observed higher carbonate alkalinity loss relative to the amount of calcium loss can be used to indicate that siderite

formed in addition to calcium carbonate. Iron (oxy)hydroxides which form in the column are converted over time

to magnetite (Odziemkowski and Gillham, 1997). Concentrations of iron are often observed to increase in the

upgradient portion of a column due to metal corrosion, followed by decreases in iron concentrations in the

downgradient part of the column as iron precipitates are formed.

Results of the influent and effluent inorganic chemical analyses can be input to a geochemical speciation model

such as MINTEQA2. In MINTEQA2, aqueous concentrations are used to calculate the saturation indices (SIs) of

various mineral species. These SIs can be used to gauge the potential for these minerals to precipitate. A negative

SI indicates undersaturation with respect to the particular mineral phase, while a positive SI indicates oversaturation.

Some caution must be used in interpreting these results, especially since such models assume equilibrium

conditions and often do not take into account the potential surface reactions involving the solid phases. Kinetic

controls on the reaction rates could, in some circumstances, result in lower rates of precipitate formation than

predicted by the speciation models.

The reliability of the column data to gauge the potential effect of precipitates on performance of field-scale PRBs

is dependent on the location of precipitation sites in-situ, and on the extent and kinetics of precipitation under field

conditions. From the calcium and alkalinity profiles collected during column tests, and the data from other field

trials, it appears that most of the carbonate precipitation occurs in the upgradient section of the reactive zones.
21

A BC

Residence Time = tC > tA > tB

Residence Time (hr)

Concentration (ug/L)

CAO

MCLB

MCLA MCLC

CBO

CCO

tA

tB

tC

t = Time

Co = Initial Concentration

MCL = Maximum Concentration Limits (Performance Criteria)

ABC = reaction pathway for compounds A, B and C

Figure 3. Conceptual residence time calulation for VOC removal.

Correction Factor

Temperature (C)

Relative Rate of Reaction

Half-Life (hr)

3.000

2.500

2.000

1.500

1.000

0.500

0.000

0 5 10 15 20 25

measured value

calculated correction factor

Figure 4. Correlation of TCE degradation rates with temperature.
22

4.2 Determination of Required Residence Time in PRBs

4.2.1 Chlorinated Volatile Organic Compounds (VOCs)

In the presence of granular iron, chlorinated organic contaminants degrade with first order kinetics. Thus required

residence times can be determined from the degradation half-life. Using half-lives and influent concentrations as

input, models are available that will calculate the time required to reach the desired treatment level. The most useful

of these models will accommodate simultaneous degradation of several compounds and will also accommodate the

simultaneous formation and degradation of breakdown products. As an example, the residence time calculation is

shown conceptually in Figure 3 for PCE degradation.

In the model, potential breakdown products are concurrently produced and degraded as described by first-order

kinetic equations. The model is an expression of the chemistry that is observed in the solution phase. For PCE, TCE,

cDCE and VC, the model takes the form

(24)

where:

f = mole fraction

k = first-order rate constant

In order to determine the VOC concentrations at a given time, first-order equations are used either directly in

commercially available software, or in their integrated form.

The residence times determined from VOC degradation rates measured in laboratory column studies of iron

materials need to be adjusted due to the effects of lower ground-water temperatures. Laboratory column studies are

normally performed at a temperature of about 20° to 25° C. Rate constants determined during laboratory studies can

be adjusted for temperature effects using the Arrhenius equation. As an example, experimental data from controlled

temperature column tests at the University of Waterloo were used to generate the linear plot of TCE degradation

rate constant vs temperature (Figure 4, Stephanie O’Hannesin, personal communication). The plot indicates that at

15° C, TCE rates could be expected to decline by a factor of 1.4 from those measured at 23° C.

There has been little published to date concerning similar relationships for other VOCs such as PCE, TCA,

cisDCE, and vinyl chloride. Limited testing at the University of Waterloo has shown little dependence of cDCE and

vinyl chloride rates on temperature. Field observations at some sites with ground-water temperatures in the order of

8°-12° C have shown apparent decreases in TCE degradation rates by factors of 2 to 2.5, but often uncertainties in

flow rates, etc., affect the interpretation of results.

4.2.2 Inorganic Constituents

Reaction rates for inorganic species vary widely, and depend on the site specific characteristics of the aquifer, the

ground water and the reactive material. Laboratory tests using site ground water, or pilot-scale field tests may be

required to accurately estimate reaction rates representative of field conditions. Estimates of reaction rates can be

drawn from batch tests, however, the accumulation of secondary reaction products may not be accurately

represented by the short-term tests. The results of column experiments may provide more accurate estimates of

reaction rates under dynamic flow conditions, but again are limited by the duration of the test.

Reaction rates drawn from laboratory experiments can be incorporated into reactive solute transport models for

estimating anticipated reaction progress within a reactive barrier. Direct application of laboratory rates may be

possible for reactions that show little temperature dependence, or dependence on site specific parameters. Reaction

rates for Cr(VI) reduction using Fe(0) shows only slight temperature dependence, allowing direct application of

laboratory rates to predict field-scale performance. Biologically-mediated systems are anticipated to be more

susceptible to variations in temperature and in nutrient concentrations. Direct transfer of laboratory-measured rates

to these systems is less certain. Use of field-scale pilot tests is warranted until a greater understanding of the

limiting factors in biologically-mediated systems is obtained.
23

Estimates of field reaction rates and determinations of residence time requirements must also include an

assessment of the potential variability of the constructed barrier. Reasonable factors of safety need to be included

in design parameters to account for construction defects and site characteristics. Ground-water flow velocities

frequently vary over small spatial intervals, resulting in sharp changes in residence times within a barrier system. In

addition, variability in the material reactivity and emplacement density may substantially influence the performance

of a barrier system (Bennett et al., 1997).

To assess residence requirements for the Elizabeth City project, five columns containing 100% granular iron and

various mixtures of iron and aquifer materials were tested using ground water from the site which contained both

TCE and Cr (VI). The column flow rates (about 2 ft/day) were similar to the flow rates predicted by a ground-water

transport model of the installed system. TCE half-lives obtained from the column tests are shown in Table 6.

Similar to other sites, TCE degraded at relatively high rates and chromium was removed quite rapidly in all

columns. The column data also indicated that VOC degradation rates rather than chromium removal rates, had the

greatest influence on the required residence time in the PRB. TCE half-lives (and the half-lives determined for

breakdown products cis-1,2-dichloroethene and vinyl chloride) were subsequently used in the field design (Bennett

et al., 1997).

4.3 Ancillary Laboratory Studies

A variety of more specialized laboratory studies may be conducted to provide additional information which can be

used in PRB design. These studies could include:

(i) laboratory hydraulic measurements (falling head tests, etc.) along with field data to provide input data

for system modeling;

(ii) surface area measurements of various candidate reactive materials. (For some contaminants, these can

be used to predict the relative performance of the respective materials);

(iii)breakdown product and reaction pathway analysis. (This could be accomplished either in batch or

column tests and may be useful in examining compounds not previously tested);

(iv) microbial analyses of solid phase or water samples from batch/column testing. (These tests may be of

particular value in evaluating PRB materials that rely on the stimulation of biological activity);

(v) tracer studies can be conducted on core samples of PRB materials to assist in evaluating the effects of

precipitate formation on system performance (Mackenzie et al., 1997; Sivavec et al., 1997).

Mineralogical study could include X-ray diffraction (XRD), to determine the presence and nature of the reaction

products, and Scanning Electron Microscopy (SEM) to determine the morphology of secondary precipitates. When

SEM is coupled with Energy Dispersive X-ray Analysis (EDX), the qualitative distribution of elements within the

reactive materials and the secondary precipitates can also be evaluated.

Other studies could incorporate surface chemical analyses, including Auger electron analyses to determine

elemental composition, X-ray photoelectron spectroscopy to determine elemental concentration and oxidation

state, and secondary ion mass spectroscopy (SIMS) to assess elemental composition and perform depth profiling of

oxidation layers.
24

5.0 Site Characterization for Permeable Reactive Barriers

A complete site characterization is of critical importance for the design and installation of a reactive barrier. This

characterization should include an evaluation of the surface features, structures and buried services to determine

whether the site is amenable to PRB installation and, if so, what types of PRB emplacement technologies are

feasible within these limitations.

The plume location and extent, ground-water flow direction and velocity, and contaminant concentrations must be

accurately known to achieve the required performance. In addition, information on stratigraphic variations in

permeability, fracturing, and aqueous geochemistry is needed for the PRB design. The plume must not pass over,

under, or around the PRB and the reactive zone must reduce the contaminant to concentration goals without rapidly

plugging with precipitates or becoming passivated. The PRB design, location, emplacement methodology, and

estimated life expectancy are based on the site characterization information, therefore insufficient or faulty

information could jeopardize the remediation effort.

In general, four aspects of site characterization should be evaluated before implementing a PRB:

• hydrogeology

• contaminant loading

• geochemistry, and when possible

• microbiology.

5.1 Hydrogeologic Characterization

As with any ground-water remediation technique, adequate hydrogeologic characterization must be done to

understand the ground-water flow patterns and the distribution of the contaminant plume. This is especially

important for the installation of a PRB since the entire plume must be directed through the reactive zone of the

barrier. To attain a “passive” remediation system, the PRB must be placed in a location that allows the plume to

move through the reactive zone under the natural ground-water gradient; i.e., the gradient must do all the work.

Information that must be obtained includes advective velocity parameters such as the piezometric surfaces (if a

confined aquifer) or gradient, the hydraulic conductivity, porosity, and the usual hydrologic parameters typical of a

careful and complete subsurface characterization. It is also important to understand seasonal changes in flow

direction and flux due to processes such as recharge, since the PRB should be designed to accommodate these

changes. An awareness of the effects of large pumping systems at the site, such as water supply wells or pumpand-treat remediation technologies, should also be known. These types of effects on flow velocity and direction can

be intermittent and have unforeseen influences on the flow entering the PRB. Figure 5 conceptually depicts how

changes in ground-water flow can reduce the effectiveness of a reactive barrier system due to incomplete plume

capture.

Beyond general hydrologic factors, the stratigraphy and lithology of the site will often dictate the type of PRB

design chosen. It is desirable to “key” the bottom of the barrier into a low-permeability clay layer to prevent

contaminant underflow, for example. If such a layer is not available at a reasonable depth, then a “hanging” design

Figure 5. Potential effects of ground water flow diversion on plume direction.

R

ec

Time

Plume and Barrier

during Rainfall Event

Plume & Barrier 

at Installation

GW Direction

New GW Direction

Due to 

Recharge

Stream Flow

Stream Flow

Recharge

Stream Stream
25

might be necessary. This type of system should be engineered to prevent contaminant underflow. In addition,

stratigraphic and lithologic features might limit the ability to perform certain installation procedures. Buried rock

might interfere with the driving of sheet piling, for example.

Understanding the vertical variation in stratigraphy is also important for choosing the stratigraphic zone(s) that the

PRB will intersect. If the contaminant plume is moving through a highly permeable layer amidst less permeable

layers, the PRB should be placed vertically to encompass this high permeability layer. It is also important that

impermeable materials, such as clays, not be “smeared” by the installation techniques across the permeable aquifer

zone that is expected to provide flow into the PRB. Therefore, a careful evaluation of the stratigraphic variability at

the location of the PRB, and the continuity of the stratigraphy with respect to the upgradient plume, will provide

confidence in the design and installation.

Stratigraphic features such as fractured rock are also important. If the plume-carrying flow zone is contacted by

zones of fractured rock, allowing some flow diversion into the fractures, then it is possible that some fraction of the

contaminant could be short-circuited around the PRB. This could give the appearance that the remediation within

the PRB is incomplete when, in fact, the PRB simply fails to intersect all the potential flow paths of the

contaminant. Reactive media emplacement into bedrock fractures may be necessary to intercept contaminant

migration pathways and achieve performance goals.

5.2 Contaminant Characterization

Information on contaminant concentrations is necessary for any successful remedial operation, whether pumpand-treat or in situ technologies are used. The remediation technique must be effective up to the maximum

concentrations and the total mass of contaminant that will be encountered at the treatment point. This is especially

important for PRBs because, once emplaced, it is difficult to change the thickness of the reactive zone. The PRB

must also be designed to eliminate the possibility that portions of the plume could flow around the barrier in any

direction. This requires a complete understanding of the extent of the plume, its width, depth, length, contaminant

concentrations within these spatial dimensions, and how these can be expected to change over time.

Because PRBs are generally placed downgradient of the plume center-of-mass, it is important that the barrier be

designed to accommodate the higher upgradient concentrations, should they arrive at the barrier unattenuated. This

requires sufficient contaminant characterization to accurately determine this high concentration zone. It is also

desirable to know whether this zone is moving downgradient over time or whether a pseudo-steady-state has been

achieved that would suggest natural attenuation is occurring. If steady-state plume boundaries have been achieved,

the PRB could be designed to transect and attenuate only the lower downgradient concentration(s) for the protection

of nearby receptors or to eliminate contaminant migration beyond site boundaries. Figure 6 illustrates the potential

result if a reactive barrier is not designed to fully handle the contaminant concentrations that reach it over time.

5.3 Geochemical Characterization

Site/plume geochemical information is needed for PRB design and implementation and to further our understanding of the expected lifetime of these systems. It has been shown, for example, that water passing through a reactive

Figure 6. Effects of plume center-of-mass downgradient, and higher concentrations impacting the barrier. Incomplete

remediation is now occurring.

Plume and Barrier as Center of

Mass Moves Downgradient

Time

Plume & Barrier 

at Installation

GW Direction

GW Direction

Stream Flow

Stream Flow

Stream Stream

Incomplete

Remediation
26

barrier of Fe(0) undergoes radical geochemical changes, including an increase in pH up to 9 or 10, elimination of

oxygen with Eh reduction to minus several hundred millivolts, and a reduction in carbonate alkalinity. Often,

sulfate is reduced to sulfide and dissolved iron appears. Other changes might also be observed, dependant upon the

contaminants being reacted and the overall ground-water geochemistry. Some of the geochemical changes can

result in precipitation onto the reactant surfaces, potentially reducing the reactivity and permeability of the reactive

zone over time. Therefore, waters high in carbonate might result in significant buildup of calcite (CaCO3

) or siderite

(FeCO3

). In addition, (oxy)hydroxides can also be expected to precipitate. Of particular potential importance is the

precipitation of ferric hydroxide, Fe(OH)3

, at the upgradient reactive zone interface due to the reaction of dissolved

oxygen with the iron. At this time, the significance of these reactions over the lifetime of a field installation is not

fully understood. However, column plugging in lab studies, particularly at the influent end where dissolved O2

contacts the Fe(0), has been observed. Laboratory research is under way to identify methods of “refreshing” the

reactivity of the Fe(0) surfaces should this become necessary.

To illustrate these possible effects, the upper graphic in Figure 7 depicts the flow of TCE-contaminated water

through a portion of a PRB of Fe(0) shortly after installation. TCE enters and ethylene exits, with a velocity of 1 and

a residence time of 1, exactly as the wall was designed. In the middle graphic precipitates have begun to

significantly coat the iron particles, limiting access of the TCE to the surface for dechlorination reactions. The

velocity is still about one, as is the residence time; however, in this system a residence time of one does not provide

sufficient contact with the reactive surfaces due to the coatings. This results in the intermediate products of

incomplete dechlorination appearing downgradient of the PRB. The bottom graphic shows that further precipitation

has blocked the pore spaces between some iron particles. As a result, the entire ground-water flux must pass through

the remaining open channels, making the velocity greater than one, the residence time less than one, and

dechlorination incomplete. This type of plugging could also result in upgradient head increase and flow around the

barrier rather than through it. It should be noted that significant plugging problems have not been observed in field

scale applications using iron to degrade TCE.

5.4 Microbial Aspects

One area of site characterization for PRBs that needs further study is microbial activity. The interactions of native

microbial populations, contaminants, and reactive barrier materials are likely to be quite complex, and have the

potential for either beneficial or detrimental effects on the remediation. Native microbial consortia are often

responsible for natural plume attenuation processes. Other beneficial effects could include enhanced contaminant

degradation within or downgradient from the PRB; for example, the further reduction in Eh due to the presence of Time

TCE

DCE

VC

DCE

VC

Ethene

Ethene TCE

TCE

TCE

residence t

vparticle

residence t

vparticle

residence t

vparticle = 1

= 1

≈ 1

≈ 1

> 1

< 1

Figure 7. Potential effects of precipitation buildup on the iron over time.
27

sulfate-reducing bacteria might increase the rates of contaminant reduction. Adverse effects could include the

potential loss of permeability of the reactive zone due to biofouling. Additional laboratory and field studies are

being conducted to understand these interactions and learn how to enhance the potential positive effects and reduce

the potential negative effects of native microorganisms.

5.5 Implementing the Field Characterization

All known information about the site should be assessed prior to mobilizing the field characterization. Usually

some information has been obtained in the process of discovering that there is a contamination problem. Due to

regulatory concerns and reporting requirements, monitoring wells may have been installed shortly after this

discovery to evaluate the contaminant(s) and its concentration. This information can be combined with historical

records, the memories of current or former workers, and surface features to obtain at least a general idea of the

location of the source term(s) and its nature. General information on topics such as aquifer locations, yields and

water table depths, water quality, stratigraphy, recharge areas, drainage basins, etc., can often be obtained from

local, state and federal agencies. Although this information might not be sufficient at the (usually) smaller scale of

the contaminated site for effecting a complete remedial design, it can be very helpful in designing the site

characterization effort.

When this information has been compiled, the choice of locations and screen depths for a few, select aquifer

testing/monitoring wells can be made, primarily to obtain general hydrologic information. Historically, plume

dimensions and site characteristics have also been determined during the installation and use of such monitoring

wells. However, these are expensive, time-consuming to install and sample, and require special multi-level

installations to provide adequate delineation in the vertical dimension. A better approach might be to begin with

surface geophysical techniques such as electromagnetic surveys, ground-penetrating radar (GPR) and seismic

studies. These techniques are of particular use when a site is completely unknown with regard to subsurface

features. Geophysical techniques can locate buried drums, pipes, and power lines in addition to providing

information on water table and bedrock depth, fractured zones, and strata thickness. Once this information is

compiled, push-tool technologies can be used for further characterization, when possible, otherwise conventional

drilling techniques may be necessary. If conventional drilling is necessary, careful consideration should be given to

the surface locations of the wells and the location of the screens with depth.

Push technologies, such as Geoprobe®, Hydropunch®, and cone penetrometers are rapidly becoming the tools of

choice for evaluating shallow plume locations in fine- to medium-grained soils of low to moderate density. Direct

push samplers can be driven rapidly and relatively inexpensively, allowing more points to be sampled than could be

accomplished with monitoring wells for the same amount of time and money. They can be used to collect soil cores,

water, and soil gas samples. Additionally, they can be used to collect these samples over very thin vertical intervals,

allowing better delineation of the contaminant concentrations in the plume and better stratigraphic characterization

than could normally be acquired with conventional exploratory drilling techniques.
28

6.0 Permeable Reactive Barrier Design

6.1 The Continuous PRB and Funnel-and-Gate Designs

The residence time required and the anticipated ground-water velocity through the PRB are used to determine the

size of PRB needed to achieve the desired treatment level. Two basic configurations are currently being used for

full-scale field application, continuous PRBs or “funnel-and-gate” designs. The effect of system configurations on

flow patterns and system dimensions can be evaluated as part of the modeling effort. Factors which may affect the

choice of configuration are discussed below.

Properly designed and constructed continuous PRBs (Figure 8a) have relatively minimal impact on the natural

ground-water flow conditions at a site. The type of continuous PRB most commonly being installed is simply a

trench that has been excavated and backfilled with granular Fe. Several other emplacement methods are possible

such as hydraulic fracturing and jetting. The continuous design incorporates no funnels and, therefore, allows the

water to pass through the barrier under its natural gradient and at its natural flow velocity. As a result, a continuous

PRB only needs to cover an area comparable to the cross-sectional area of the plume. The ground-water flow

velocity through the PRB will be very similar to the velocity in the aquifer. As long as the hydraulic conductivity of

the aquifer is less than that of the PRB, underflow of contaminated ground water should not occur.

Ideally the continuous PRB is built to a depth that somewhat over-encompasses the vertical and horizontal

dimensions of the contaminant plume, as a safety factor, and is filled with granular iron or some other reactive

material. The PRB thickness must be sufficient to remediate the contaminant of concern to the established

concentration goals. As with the funnel-and-gate design, it is desirable to place the bottom of the continuous PRB

into an impermeable zone; i.e., “key” it into impermeable strata to mitigate the potential for contaminant underflow.

The upgradient and downgradient surface areas of the aquifer material contacting the PRB will be approximately

the same, minimizing disruption in the natural ground-water flow relative to current funnel-and-gate designs

(Figure 8b).

GW Flow

Continuous PRB

Contaminant Plume

GW Flow

Reactive Gate

Funnel-and-Gate PRB

Funnel

Contaminant Plume

Figure 8a. Plume capture by a continuous PRB trenched system. The plume moves unimpeded through the reactive zone.

Figure 8b. Plume capture by a funnel-and-gate system. Sheet piling funnels direct the plume through the reactive gate.
29

A difficulty with the trench emplacement involves trenching and filling in the fully saturated zone due to the

problems of immediate water intrusion and potential collapse of the trench walls. This problem requires that the

aquifer materials be removed and the reactive media emplaced nearly simultaneously unless subsurface constructions (sheet pile walls, trench filling with biodegradable slurries such as guar, etc.) are used. This rapid aquifer

material removal/reactive media emplacement is possible using continuous trenching devices such as the one shown

in Figure 9. During June 1996 a 150 foot long, 2 foot wide, 24 foot deep continuous trench PRB of iron was

installed at the U. S. Coast Guard Air Support Center near Elizabeth City, North Carolina, to intercept a mixed

plume of chromate and TCE, following the success of a small-scale field test (Puls et al., 1996). This marked the

first time that a continuous trenching technique had been used to emplace iron in a subsurface reactive barrier. A

drawback of this type of PRB installation is the difficulty tracking the volume of reactive material actually

emplaced during the construction.

In a funnel-and-gate configuration, low permeability funnels direct ground water toward permeable treatment

zones or gates.The basic design of a funnel-and-gate system is shown in Figure 8b. The “funnel” typically consists

of sheet pilings, slurry walls, or some other material and is preferably “keyed” into an impermeable layer (clay,

bedrock) to prevent contaminant underflow. This funnel is emplaced to encompass and direct the flow of

contaminated water to a “gate” or “gates” containing a permeable zone of granular Fe(0) or other reactive material.

Due to directing a large cross-sectional area of water through the much smaller cross-sectional area of the gate,

ground-water velocities within the gate will be higher than those resulting from the natural gradient. The funnel

portion of the design is engineered to completely encompass the path of the contaminant plume and the overall

design must prevent the contaminant plume from flowing around the barrier in any direction. The gate shape may

be controlled by construction techniques, but have most commonly been rectangular. Recent variations include the

use of backfilled caissons, media filled hollow vibrating beams, or emplaced reaction vessels. For the emplaced

reaction vessels, contaminated upgradient waters are directed into the subsurface vessel which contains reactive

media, and the treated water is discharged through a pipe that extends downgradient through the impermeable wall.

The more typical rectangular, or box-shaped gate can be built by driving temporary sheet pilings and/or building

removable subsurface walls within which the reactive materials are emplaced. Sheet pilings can be driven to

delineate the sides and ends of the media gate. The interior of this construction can then be dewatered and excavated

to make room for the reactive material (Gillham and Burris, 1992). Water intrusion is reduced by the impermeable

pilings and any seepage can be pumped away. This construction approach also allows the incorporation of other

features, such as additional treatment zones or monitoring zones.

The permeability of the reactive material in the gated zone must be equal to or greater than the aquifer

permeability to minimize flow restrictions. At the same time, it is necessary for the flowpath length through the gate

to be long enough for complete contaminant remediation; i.e., sufficient contact, or residence time, must be allowed

between the reactive material and the contaminant. Overall system length depends on the number, location and size

of treatment gates and should be determined through ground-water flow modeling. Particular care is required in

designing and constructing the connection between the impermeable “funnel” section and the permeable “gate”

section in order to avoid bypass of contaminated ground water.

Some problems can accompany the funnel-and-gate design that should be addressed, prior to installation, with a

careful site characterization. Foremost among these is the potential for diversion of ground water around the funneland-gate. Even if the permeability of the gate section is tremendously enhanced relative to the aquifer permeability,

the gate again contacts the same aquifer material on the downgradient side and permeability is immediately reduced

to the initial aquifer value. It is important that the gated zone and the downgradient gate/aquifer interface be able to

pass the flux of contaminated ground water intercepted by the funnels, if not the entire flux of intercepted water

(both contaminated and uncontaminated). It would seem unlikely that the total volume of water in a high flux

system, directed by a high surface area funnel, could be infiltrated through the much smaller surface area of the gate

and aquifer material interface that contacts the downgradient side of the gate. This constriction of flow could result

in a significant head difference across the reactive barrier. In fact, a seven to ten foot head difference reportedly

developed across the funnel and multiple-gate system that was installed at the Denver Federal Facility, Denver, CO

(Cushing et al., 1996) and has continued to persist (McMahon, 1997). The important point when designing a funneland-gate system, however, is not that some water is diverted around the barrier, but that none of the contaminated

water is diverted and it passes through the reactive zone for treatment.

Several combinations or “treatment trains” of different types of reactive materials might be considered at specific

sites containing a combination of contaminants (e.g., metals and VOCs, or VOCs and petroleum hydrocarbons).

The funnel-and-gate PRB configuration may be more appropriate for these systems, providing a “focused”
30

treatment area. This may be of particular value if the materials need to be replaced or replenished periodically over

the life of the system. These “focused” treatment areas will also facilitate performance monitoring of the system.

However, research needs to be done on how to integrate differing reaction technologies into a subsurface “treatment

train,” so that the reactions do not interfere or limit one another.

Many of the reaction mechanisms associated with PRB materials are surface area dependent. The amount of

reactive material needed is, therefore, proportional to the mass flux of contaminant requiring treatment. Therefore,

both the funnel-and-gate and continuous PRB configurations should theoretically require the same amount of soil

excavation and disposal since the PRB should contain the same volume of reactive material. However, in an actual

field implementation, it is possible that more iron will be required by a continuous PRB using current commercialscale installation techniques (trenching), particularly if the plume is very broad and requires a relatively long PRB.

The reason for this is the need to be certain that the PRB has no unreactive gaps or flowpaths throughout its length.

There will be some limiting thickness that has to be maintained to assure PRB integrity throughout its volume. The

use of jetting, as an emplacement method rather than trenching, could potentially mitigate this problem for thin

continuous PRBs.

Given the above, the selection of an appropriate configuration should be made on a site-specific basis. The cost of

construction of either configuration will ultimately be dependent on the depth, width and saturated thickness of the

plume, which controls the overall dimension of the system. Costs for reactive materials, “funnel” materials (if

required), and construction will be the major capital cost components of these systems.

When considering emplacement methods and system configuration, several other factors common to subsurface

construction procedures need to be considered, including:

• the need for dewatering during excavation;

• the means and costs of ground water and soil disposal;

• health and safety. The only “hazard” associated with granular iron used to date has been nuisance dust.

However, the entry of construction personnel into gate excavations does create a health and safety issue;

• disruption to site activities

These factors will also influence the cost of the system.

6.2 Emplacement Methods and Comparisons

The installation methods used thus far for PRBs are few in number. Most systems have relied on standard means

of excavating and backfilling to place reactive material in the path of a plume, as was discussed for the continuous

PRB and the funnel-and-gate designs. Likewise, the impermeable walls of funnel-and-gate systems have used sheet

pile driving or slurry wall constructions. Among the other methods being developed, tested, and implemented are

the use of trenching machines, mandrel or tremie tubes, deep soil mixing, high-pressure jetting and vertical

hydraulic fracturing. It is unlikely that any single approach will be found to be feasible and cost-effective for all

types of applications. Rather, the value of having a variety of methods from which to choose will be the ability to

select the best method for each application, with its particular combination of depth requirement, reactive media

thickness requirements, stratigraphy, sensitivity to spoils generation, and other factors. Calculations based on the

equations in AppendixC can be used to estimate whether a given emplacement method can provide sufficient iron

thickness to accomplish site-specific remediation goals, provided the method also achieves the other emplacement

requirements.

6.2.1 Conventional Excavation

Conventional excavation is a basic construction practice using equipment such as backhoes, excavators, and

cranes. Standard excavators can reach depths of about 35 feet, and modified excavators can reach depths of about

70 feet. Cranes fitted with clamshells can attain greater depths.

Shallow trenches may not require any support to remain open before backfilling, but most PRB applications must

be deep enough that they will need to be supported prior to being backfilled with either permeable or impermeable

materials. There are two means of supporting trenches: structural walls (e.g., trench boxes) and slurries.

The trench box approach has been used to install the gate portions for most PRBs thus far. Often the final stages

of excavation involve time-consuming confined entry procedures, making gate installation the most expensive

component of a PRB project. As a result, some recent projects have tested other means of supporting trenches,
31

namely caissons (used for the Somersworth, NH, and Dover Air Force Base projects), continuous trenching

machines (used at Elizabeth City, NC, Oregon, New York and California) and biodegradable slurries (Vancouver,

Canada).

The use of slurries to support trenches prior to backfilling may provide significant cost savings. Bentonite slurries

are commonly used for this purpose in constructing slurry walls. Bentonite slurries cannot be used to construct

permeable zones, but using a biodegradable polymer solution such as guar gum may be feasible and cost-effective.

A biodegradable slurry was used to install a small-scale test PRB (30ft x 6ft x 30ft) in Vancouver, Canada. Tests are

ongoing to evaluate the effects (if any) of the biodegradable slurries on iron and other reactive materials.

6.2.2 Trenching Machines

Trenching machines are an effective means of installing relatively shallow trenches. The type most applicable to

installing PRBs are of the “chain saw” design, which cuts a trench between 12 and 36 inches wide. During

installation, the trench is held open by a trench box attached to the chain saw cutting belt mechanism. The trench

box is pulled along after the cutter chain. A hopper attached to the top of the trench box can be filled with reactive

material and feeds the material into the trench. Typical trenchers are limited to depths of about 20 to 30 feet, but

larger equipment is expected to be more widely available in the future. Based on calculations presented in

Appendix C, for a trench width of 36 inches and a granular iron bulk density of 160 lb/ft2

, the maximum weight-perarea achievable by this method is 480 lb/ft2

.

6.2.3 Tremie Tube/Mandrel

A tremie tube approach is a modification of the means of installing wick drains. A hollow rectangular tube with

expendable drive shoe on the bottom is driven to depth with hydrostatic force or a vibratory hammer. The tube can

then be filled with dry granular material or a slurry containing the reactive media. The tube then is extracted,

leaving the drive shoe and added materials in the ground. Then the process is repeated along the desired path, each

emplacement overlapping the previous one by an amount necessary to provide continuity.

This method was used to install iron-containing zones for the Phase IIA LasagnaTM project at the DOE gaseous

diffusion plant in Paducah, Kentucky. In that case, mixtures of iron/coke or iron/clay were installed to a depth of

45 feet. At a bulk density of 160 lb/ft3

, the use of 100% granular iron in a 4-inch zone would give a weight-per-area

of slightly more than 50 lb/ft2

 (see Appendix C).

6.2.4 Deep Soil Mixing

Deep soil mixing utilizes large augers (e.g., 3 ft to 8 ft diameter) suspended by cranes and driven by large motors.

Once a zone is mixed, the auger and crane are indexed to the next location. Applications carried out to date include

injecting cement or other grouting agents to build containment walls or to stabilize an area. In addition, deep soil

mixing has been used with air or steam injection to volatilize contaminants.

It is thought that relatively large amounts of iron can be emplaced by deep soil mixing, but little data is available.

Assuming that the resulting zone is 50% iron, 3-foot augers could be used to apply about 180 lb/ft2

, and 8-foot

augers could apply up to about 480 lb/ft2 (see Appendix C).

An important aspect of deep soil mixing is the relatively high mobilization costs associated with transporting and

setting up the large equipment involved. For small projects, this may be prohibitive.

6.2.5 High-Pressure Jetting

High-pressure jetting is an established practice to inject grouting agents to improve the structural characteristics of

soil for construction purposes. More recently, it has been used to inject grouts to make impermeable walls. Jetting

is an attractive concept for many reasons. An important advantage in certain settings will be the ability to install a

wall around obstructions such as boulders and utility lines. Also, the equipment is small and the mobilization costs

are relatively low.

This approach uses jetting nozzles incorporated into a specialized section of the drill string located above the drill

bit. Once the drill string reaches the desired depth, the pump increases its output (up to 90 gpm and 6,000 psi). As

the slurry is injected into the formation, the drill string is extracted from the borehole at the desired rate. Due to the

jetting process, most of the finer soil fractions are forced to the surface, but a significant portion of the coarser

fractions remain.

If the jetting nozzle is rotated during extraction, a column of injected material forms which is approximately three

to seven feet in diameter. Depending on the pumping and extraction rates, it is anticipated that the columnar zone
32

will contain up to 75% of the injected reactive material. Therefore, a 3-foot diameter permeable reactive zone might

contain up to 360 lb/ft2

 granular iron (see Appendix C).

If the jetting nozzle is not rotated during extraction, it creates an injected zone known as a thin diaphragm wall. If

such a wall averages five inches thick and contains 75% iron, then the resulting weight-per-area is 50 lb/ft2 (see

Appendix C).

A pilot test to inject granular iron was conducted recently by DuPont. Guar gum (a natural food thickener) was

used as the viscosifying agent for making slurries of 50-mesh iron particles. Both columnar and thin diaphragm wall

injections were successfully demonstrated. The depth of injection was only fifteen feet in this case, but the principle

of using high pressure jetting was established, and it appears to be ready for use at greater depths.

6.2.6 Vertical Hydraulic Fracturing and Reactant Sand-Fracturing

In vertical hydraulic fracturing (VHF), holes are bored to initiate a fracture in permeable sands. A Fe(0)-

containing fracturing fluid gel is then pumped into the fractures, forming a continuous wall of reactive material. The

gel can be guar gum, which dissolves and leaves a PRB of Fe(0). The fracturing fluid used at the Caldwell trucking

site, New Jersey, consisted of potable water, guar gum, a borax cross-linker (to link iron to the gel), pH buffer, an

enzyme breaker (to break down starch in the guar after injection), and a fine-sand propant (Hocking, 1998). Crosslinking was necessary to prevent the iron from falling from the gel before permeating the fracture. Because many

undesired reactions can occur between gel ingredients, potentially reducing the reactivity of the iron, proposed

fracture fluid ingredients should be extensively tested prior to using this installation method.

Reactant sand-fracturing (RSF) uses high-pressure fracturing with a sand propant, taking advantage of the

fractures that exist in bedrock, and providing a means of creating reactive fracture zones of Fe(0) within

contaminated bedrock. As with VHF, a reactive fracturing fluid is needed for RSF since granular iron and iron

filings do not have the needed hydraulic properties. An iron foam propant was chosen for a pilot test at a California

facility as a method for placing reactive Fe(0) media into a fractured bedrock aquifer contaminated with chlorinated

solvents and metals (Marcus, 1998). The pilot test consisted of boring holes, with chemical and physical testing to

identify fractures containing contamination, pretreatment hydraulic fracturing (to ascertain that the injection

equipment could fracture the bedrock), injection of the foamed iron propants, and post-treatment confirmation of

emplacement with down-hole geophysical and geochemical testing. Cross-linkers in the propant were not necessary

due to the low density of the iron foam relative to some other Fe(0) forms. The low-density foam also allowed the

use of standard hydraulic fracturing equipment.

6.3 Emplacement Verification

As part of the quality assurance program for the installed system, certain measures should be used to verify

emplacement of the PRB system as designed. This is most critical for continuous barrier configurations, but also

required for funnel-and-gate and other emplacement designs. Whether the PRB is installed with excavation and fill,

continuous trenching, hydrofracturing, or jetting, its emplacement and integrity should be confirmed. This will

entail additional analyses beyond contaminant disappearance in downgradient monitoring wells. Geophysical

methods have been used to a limited extent in this regard. These methods include natural gamma, conductivity,

electrical resistivity, cross borehole tomography and surface radar. Tracer tests and standard hydrologic characterization methods can also provide useful information, especially when the data are combined with and compared to

the data from the geophysical studies.
33

7.0 Monitoring Permeable Reactive Barriers

7.1 Planning the Monitoring Effort

A Quality Assurance Project Plan (QAPP) must be developed prior to any ground-water sampling, including

baseline water quality data collection efforts. The QAPP must address the data quality requirements established for

the project. U.S EPA’s Pocket Guide for the Preparation of a Quality Assurance Plan (EPA/600/9-89/087) is a

useful resource in this regard. A monitoring plan is an integral part of the QAPP for the project and emphasis should

be placed on data collection methods and monitoring network design equal to that traditionally placed on analytical

methods and sample handling.

All procedures and techniques used for site characterization, ground-water monitoring well installation and

development, sample collection, sample preservation and shipment, analytical procedures, and chain-of-custody

control should be specified in a QAPP. At a minimum, the QAPP should address

• sampling objectives

• pre-sampling activities

• sample collection

• in-situ or field analyses and equipment calibration

• sample preservation and handling

• equipment decontamination

• chain-of-custody control and records management

• analytical procedures and quantification limits for both laboratory and field methods

• field and laboratory quality assurance/quality control

• evaluation of data quality, and

• health and safety.

Other sources of useful information on the topics of QA procedures and ground-water monitoring include the

following: Chapter Eleven “Ground Water Monitoring” of EPA’s manual titled Test Methods for Evaluating Solid

Waste (U.S. EPA 1986c, commonly known as “SW-846”); Ground Water Sampling - A Workshop Summary

(EPA/600/R-94/205, 1995); Subsurface Characterization and Monitoring Techniques - A Desk Reference Guide,

Vol. 1: Solids and Ground Water and Vol. 2: The Vadose Zone, Field Screening and Analytical Methods

(EPA/R-93/003b, 1993).

7.2 Compliance Monitoring

7.2.1 Objectives

Compliance monitoring typically involves the monitoring of the contaminants of interest at a particular hazardous

waste site where dissolved concentrations have been detected that exceed regulatory limits. The focus is on the site

and its compliance points with the monitoring program driven by regulatory requirements. General water quality

monitoring is also often included, such as determinations of major cations and anions and other water quality

indicator parameters such as pH, alkalinity, specific conductance, etc. For permeable reactive barriers, similar

compliance monitoring requirements are necessary; however, the placement and design of monitoring wells or

points and the methods used to sample ground water may be different. In addition, monitoring of degradation

products from reductive dehalogenation reactions or other contaminant transformation products may be required.

7.2.2 Compliance Monitoring System Design

Well placement and design are important to ensure adequate assessment of system performance. In addition to

upgradient and downgradient wells, wells should be located to ensure that contaminated water is not flowing

around, under, or over the barrier wall. The number of wells used will depend on system design and size. Figure 8

shows two examples for a funnel-and-gate design and a continuous wall design. Use of two inch diameter wells is

usually sufficient and appropriate for compliance monitoring purposes. Smaller diameter wells or piezometers can

be used if the selected sampling devices permit it. Selection of screen length should be compatible with sampling

program objectives and site conditions, particularly with respect to plume location as it exists in three dimensions

within the aquifer (Powell & Puls, 1997c) and the depth of the reactive zone. Compliance wells located near the

permeable reaction zone should be located far enough away to avoid mixing of water from distinct geochemical
34

zones during purging and sampling of the wells. For example, a well located immediately downgradient of the PRB

should be located such that water is not extracted from the reaction zone itself during well sampling. However, the

well should not be located so far downgradient that it takes a long time to determine the effectiveness of the PRB.

Similarly, wells located upgradient should not be too close to the PRB, such that water is inadvertently pulled from

the reaction zone during purging and sampling operations. Wells located beneath the PRB should also be carefully

located with respect to the target sampling zone. The sampling device and sampling methods employed will dictate

how close wells can be located to the PRB and some thought must be given to the zone of influence during the

sample collection process.

It should also be realized that if the PRB has been located to transect the contaminant plume rather than being

installed entirely downgradient of the plume, compliance wells downgradient of the PRB will probably not show

immediate reductions in contaminants. This is because diffusion and desorption of the contaminant from the aquifer

materials downgradient of the PRB will continue for some time before gradually diminishing. Well number 46 at

the Elizabeth City, North Carolina, PRB site is a good example of this phenomenon. The PRB was installed in June,

1996, with Well 46 (Figure 9) at a distance of about 56 feet downgradient of the PRB. In November of 1996, five

months after PRB installation, the TCE concentration at Well 46 was 256 μg L-1. The following September (1997),

the TCE concentration was approaching compliance for TCE (10.9 μg L-1 ) and was in compliance for VC.

7.2.3 Compliance Sampling Methods

Low-flow sampling methods are recommended for compliance sampling purposes (Puls and Barcelona, 1996).

Low-flow refers to the velocity with which water is withdrawn from the aquifer and is usually from 0.1 to

0.5 L/min, depending upon site-specific hydrogeologic conditions. The effectiveness of using low-flow purging is

intimately linked with proper screen location, screen length, and well construction and development techniques.

The objective is to pump in a manner that minimizes stress to the hydrogeologic system, to the extent practical,

taking into account established program sampling objectives.

The reestablishment of natural flowpath equilibrium in both the vertical and horizontal directions following

disturbance of the water in the well and prior to sampling are important for correct interpretation of the data (Powell

and Puls, 1993). Much of the need for purging has been found to be due to passing the sampling device through the

overlying casing water, which causes mixing of these stagnant waters and the dynamic waters within the screened

interval. These disturbances and impacts can be avoided using dedicated sampling equipment set within the

screened interval, which precludes the need to insert the sampling device prior to purging and sampling (Powell and

Puls, 1997c).

Figure 9. Continuous trencher used at Elizabeth City, North Carolina.
35

For high resolution sampling needs, screens less than 1 m in length should be used. Isolation of the screened

interval water from the overlying stagnant casing water may be accomplished using low-flow minimal drawdown

techniques. If the pump intake is located within the screened interval most, if not all, of the water pumped will be

drawn directly from the formation with little mixing of casing water or disturbance to the sampling zone. However,

if the wells are not constructed and developed properly, zones other than those intended may be sampled. At some

sites where geologic heterogeneities are sufficiently different across the screened interval, higher conductivity

zones may be preferentially sampled. This is another reason to use shorter screened intervals, especially where high

vertical spatial resolution is a sampling objective.

It is recommended that water quality indicator parameters be used to determine purging needs prior to sample

collection in each well. Stabilization of parameters such as pH, specific conductance, dissolved oxygen, oxidationreduction potential, temperature and turbidity should be used to determine when formation water is accessed during

purging. In general, pH stabilizes first, along with temperature and specific conductance, followed by oxidationreduction potential (redox), dissolved oxygen and turbidity (Puls et al., 1992; Puls and Powell, 1992a). Temperature

and pH, while commonly used as purging indicators, are actually quite insensitive in distinguishing between

formation water and stagnant casing water; nevertheless, these are important parameters for data interpretation

purposes and should also be measured. Instruments are available which utilize in-line flow cells to continuously

measure the above parameters. Most of these same parameters are important performance measures for reactive

barriers (e.g., for iron corrosion reactions occurring within the treatment zone).

It is important to establish specific well parameter stabilization criteria and then consistently follow the same

methods thereafter. Generally, the time or purge volume required for parameter stabilization is independent of well

depth or well volumes for low-flow sampling. Important variables are well diameter, sampling device,

hydrogeochemistry, pump flow rate, and whether the devices are used in a portable or dedicated manner. If the

sampling device is already in place (i.e., dedicated sampling systems), then the time and purge volume needed for

stabilization is much shorter. Other advantages of dedicated equipment include less purge water for waste disposal,

much less decontamination of equipment, less time spent in preparation for sampling as well as time in the field, and

more consistency in the sampling approach which will probably translate into less variability between sampling

results. The use of dedicated equipment is strongly recommended at wells which will undergo routine sampling

over time, as will be the case with monitoring of permeable reactive barriers. The use of low-flow sampling

techniques will also allow the placement of compliance wells closer to the treatment wall system, since the reduced

purge volume will decrease the chance of mixing waters from unintended sources and zones.

 Monitoring well sampling should always proceed from the well that is expected to be least contaminated to the

well that is expected to be most contaminated, to minimize the potential for cross-contamination of samples that

may result from inadequate decontamination of sampling equipment. Samples should be collected and containerized according to the volatility of the target analytes. The preferred collection order for some of the more common

ground-water analytes is as follows (Barcelona et al., 1985):

1. Volatile organics (VOAs or VOCs);

2. Dissolved gases and total organic carbon (TOC);

3. Semivolatile organics (SMVs or SVOCs);

4. Metals and cyanide;

5. Major water quality cations and anions;

6. Radionuclides.

7.2.4 Compliance Sampling Frequency

The frequency with which compliance samples are collected should be based on the hydrogeologic character of

the aquifer, the proximity of sensitive receptors such as water supply wells, and the risk posed by the contaminant(s).

In general, if ground-water velocities are low, then samples can be collected less frequently than when velocities are

high. As a general guideline, quarterly monitoring seems appropriate, at least initially, for most sites, with

variations permitted for exceptionally slow or fast ground-water flow situations. In many cases, when the groundwater flow rate is very slow, semiannual sampling may be justifiable.

Due to the massive subsurface disturbance created during the construction and installation of a permeable reactive

barrier, initial sampling results are often not representative of system performance. Re-equilibration of the groundwater flow field is required before sampling results can be properly interpreted. Based on field results thus far, this

seems to take about 1-3 months. Water level measurements should be taken within the first few days, both
36

upgradient and downgradient of the PRB, and continued on a regular basis to ascertain that the system is

functioning as designed with respect to plume capture, and the establishment and maintenance of designed flow

parameters through the system and upgradient of the treatment zone. These data are relatively easy and inexpensive

to obtain and the frequency of collection can therefore be increased compared to samples requiring expensive

analysis.

The compliance sampling program should be reevaluated and potentially revised 1-2 years after the installation

and initial data review. The sampling frequency may be reduced if quarterly sampling shows consistent results and

other performance parameters indicate that the system continues to perform as designed.

7.2.5 Contaminant Breakthrough/Bypass and Formation of Undesirable Products

The primary objective of the compliance sampling program is to determine whether the treatment wall is meeting

design goals for remediating the contaminated ground water. The presence of contaminants which exceed target

cleanup goals in downgradient water samples is the first compliance measure most people, particularly regulators,

will examine. It is equally important to examine other data with regard to meeting compliance goals. For example,

hydraulic head data is arguably of equal importance to contaminant concentration data for compliance purposes.

Changes in water flow patterns may lead to short circuiting of the PRB by the plume. The plume could dip below

or move around the barrier due to changes in hydraulic head which could drastically alter previous flow paths.

Underflow is a concern not only for hanging wall configurations (Figure 1), but also for designs keyed into lowpermeability layers. Also, it should not be assumed that “impermeable” layers are perfectly flat, continuous, or even

impermeable. There is typically never enough site characterization data to provide this assurance, so careful

compliance monitoring is important.

It is also important to thoroughly understand the reactions which cause the transformation and destruction/

immobilization of the contaminants to be able to monitor for undesirable degradation or transformation products as

part of the compliance sampling program. As was shown earlier for trichloroethylene (TCE) and Fe(0), there are

two potential pathways for the reductive dechlorination to ethene and ethane to occur: sequential hydrogenolysis,

and reductive ß- elimination, each leading to ethene and ethane as final products (Roberts et al., 1996). Possible

intermediate products are cis- 1,2-DCE and vinyl chloride (VC). Monitoring for these intermediate constituents

should be included as part of the compliance sampling program since they are also regulated toxic compounds.

In the case of chromium and Fe(0), the reaction product is an insoluble hydroxide mineral phase. This can only be

confirmed using advanced surface analytical techniques, but can be inferred from non-detection of Cr(VI) and

minute or non-detection of Cr(III) in aqueous samples together with ground-water quality data and geochemical

modeling.

Evaluating trends in the data over time is important since this technology is new and many questions remain

concerning long-term in situ performance, especially with regard to maintenance of reactivity and system

permeability.

7.2.6 General Water Quality Parameters

Water quality parameters such as pH, specific conductance, alkalinity, major cations and anions and dissolved

oxygen are routinely collected as part of many site assessments and remedial actions. These same data should be

collected as part of the compliance sampling program for permeable reactive barriers. Also, parameters relevant to

the reactive material used in the PRB should be considered for analysis. For example, when zero-valent iron is the

reactive material, measurements such as redox (Eh) should be made and samples analyzed for dissolved ferrous

iron. All of these parameters are important as either indicators of the corrosion process for iron-mediated

transformation reactions, or for potentially undesirable secondary effects in downgradient water quality. Most of

the effects due to the emplacement of a reactive iron barrier should disappear within a few meters downgradient of

the reaction zone, due to the buffering of subsurface systems, but changes in downgradient water quality should be

followed as part of the routine compliance monitoring program for this new and innovative remedial technology.

Physically or chemically unstable analytes should be measured in the field, rather than in the laboratory. Examples

of unstable parameters include pH, redox potential, dissolved oxygen, alkalinity, temperature and dissolved iron.

Dissolved oxygen (and other analyses that would be expected to vary significantly with dissolved oxygen or carbon

dioxide concentration), turbidity, and specific conductance should be determined in the field as soon as practicable

after purging. Most conductivity instruments require temperature compensation; therefore, the temperature of the

samples should be measured at the time conductivity is determined unless the monitoring equipment automatically

makes this compensation. Temperature data should also be collected as near the wellhead as possible. This reduces

travel distance through the tubing and minimizes changes due to the atmospheric temperature.
37

Three methods generally are employed for measuring unstable field parameters: use of an in-line flow cell,

collecting discrete samples and analyzing them at land surface, and using analytical equipment with probes that can

be lowered into the well. The preferred method is use of an in-line flow cell. Although some investigators have

experienced freezing of tubing in very cold weather, this method provides results that typically are more precise

than those obtained using down-hole probes or collecting discrete samples and analyzing them at land surface.

Analyzing discrete samples at the land surface involves collecting a sample in a clean bottle or beaker in the same

manner that a sample for laboratory analysis would be collected, and then analyzing the sample using a field test kit

or meter. Use of down-hole probes typically requires that investigators induce ground-water movement past the

probe, either by physically moving the probe (potentially creating turbidity and potentially mixing casing water

with screened interval water), or pumping (potentially producing inconsistent results from well to well). If downhole probes (e.g., pH electrode, thermistor) are used to measure unstable parameters, the probes should be

decontaminated in a manner that prevents them from contaminating the water in the well. In no case should field

analyses be performed directly on samples that will be submitted for laboratory analysis.

7.3 Performance Monitoring

7.3.1 Objectives

Performance monitoring is focused on the PRB system itself (including impermeable funnels, if present), rather

than the entire site or the compliance boundaries. Effective performance monitoring begins with adequate site

characterization, to provide a baseline for later comparison, and its objective is to evaluate PRB performance

relative to design. Performance monitoring of PRBs includes the evaluation of physical, chemical and mineralogic

parameters over time. It should address verification of emplacement and be able to detect loss of reactivity, decrease

in permeability, decrease in contaminant residence time in the reaction zone, and short circuiting or leakage in the

funnel walls. In addition to monitoring the contaminants of concern and general water quality, the following are

also recommended: contaminant degradation products, precipitates, hydrologic parameters and geochemical

indicator parameters. Understanding the mechanisms controlling contaminant transformation, destruction or

immobilization within the reaction zone is critical to interpretation of performance monitoring data as these data

provide insight into barrier functionality.

7.3.2 Performance Monitoring System Design

Sample points to evaluate system performance are located within or immediately adjacent to the PRB system. Due

to the relatively small zones within the aquifer where the reactive material is emplaced, these sampling points and

the sampling methods can be somewhat different from the more traditional approaches employed for compliance

monitoring purposes. Also, the volume of aquifer targeted for performance sample collection is usually significantly smaller than that targeted for compliance monitoring purposes.

Monitoring wells can be inserted into the reactive barrier zone easily and precisely if a shallow system is being

constructed in a funnel-and-gate design. In some funnel-and-gate installations this has been accomplished by

suspending the wells in a metal framework which is subsequently removed during backfilling of the treatment

material. In a continuous wall design, where a trencher continuously excavates aquifer sediment and backfills with

the reactive material in one step, this is not an option. Monitoring points in such designs must be installed later. This

can be done using small drill rigs or by hydraulic push to place the well points. Precise placement of well points

within the barrier can be problematic, particularly with increasing depth and with thin PRBs. Small vertical

deviations can quickly result in screen placement into unintended zones. For example, in a 20 inch wide PRB where

a well is positioned in the center of the barrier, a 3% deviation from vertical will cause the well point to exit the

reactive media only 16 feet below ground surface. There are, however, rather simple and effective means to verify

proper placement. Certain geochemical indicator parameters can confirm emplacement within zero-valent iron

barriers. The use of a simple conductivity probe (commercially available for use in hydraulic push-type soil probes)

can also verify iron placement because conductivity readings between the iron and the aquifer sediment can differ

by as much as two orders of magnitude.

Smaller diameter wells (e.g., 3/4 inch i.d.) or even bundled tubes with short screens (< 1 ft screens) are often

preferred for performance monitoring purposes (Figure 10). Proper location of the screened intervals in three

dimensions is important for monitoring different vertical zones within the PRB as well as to provide coverage for

different flow paths and allow estimation of degradation or transformation rates as the contaminants pass through

the system. Similarly, these same monitoring points can be used for tracer tests to evaluate changes in permeability.

The placement of monitoring points for flow path analysis requires prior detailed site characterization to determine

flow path direction, flow velocities and system heterogeneities.
38

7.3.3 Performance Sampling Methods

Monitoring of permeable reactive barriers presents many unique challenges. Traditional sampling approaches are

often inappropriate because the withdrawal of large water volumes might compromise PRB sampling objectives.

Data is often sought to confirm not only contaminant reduction or removal but to provide data on contaminant

transformation or degradation rates and changes in the wall reactivity. These rates and changes might need to be

addressed over relatively small spatial intervals. To do this accurately, water from much smaller volumes of the

aquifer or from within the PRB must be withdrawn.

The use of passive or semi-passive sampling approaches can provide the required samples for these objectives

(Powell and Puls 1997c). Several researchers have demonstrated the effectiveness of using discrete-level sampling

approaches (Robin and Gillham, 1987; Powell and Puls, 1993). Puls and Paul (1995) have shown the effectiveness

of a discrete multi-level sampling device which can be used in traditional monitoring wells to provide no-purge

samples representative of formation water. However, this work addressed only inorganic contaminants and

inorganic constituents in ground water. Researchers at NRMRL-EPA are currently evaluating the use of this device

for chlorinated organics and to understand the effects that highly reducing environments might have on sampling

results.

7.3.4 Performance Sampling Frequency

The frequency for performance sampling will be dictated by site-specific hydrogeochemical conditions, system

design, and performance sampling objectives, which should be specified prior to system installation. These

objectives should be discussed among the responsible parties, the lead regulatory agency and those who will be

tasked with implementing the sampling program and maintaining the remedial system. At current installations the

scope of performance sampling activities ranges from almost no monitoring to extensive and detailed performance

monitoring. Many of the heavily monitored installations are being used as research sites to resolve questions

concerning effective implementation of the technology and long-term performance issues. As these questions are

answered and issues are resolved, it is anticipated that subsequent performance sampling events will be greatly

reduced over what is currently practiced. However, there will always be a need for some level of performance

sampling (or at least the ability to perform such sampling on a contingency basis, should unexpected compliance

results occur) to insure that the system continues to operate as designed, particularly with respect to reactivity and

permeability.

Figure 10. Bundled tubes with short screens for performance monitoring.

8

7

6

5

4

3

2

1

0

1/2" PVC wells with

4" screens

2" PVC well with

10' screen

Distance Below

Ground Surface

(meters)
39

7.3.5 Contaminant Degradation/Transformation

As with compliance monitoring, it is important to evaluate or confirm that the desired degradation or transformation of target contaminant species is occurring as the plume moves through the reaction zone. In addition to those

degradation products which may also be regulated contaminants, it may be desirable to analyze for non-contaminant

species which provide assurance of contaminant transformation.

7.3.6 Geochemical Indicator Parameters

Geochemical indicator parameters which provide some measure of system performance will vary based on the

reactive barrier material and the contaminants present. For a reactive iron barrier these will include: pH, Eh,

alkalinity, dissolved oxygen, total dissolved sulfide, ferrous iron and, potentially, dissolved hydrogen. These

parameters can indicate that iron corrosion is proceeding and provide some indication of the extent of precipitate

formation within the barrier that may eventually decrease wall performance over time. When included with general

water quality monitoring and used in conjunction with geochemical modeling, these geochemical parameters can

support modeling projections concerning potential precipitate formation.

For example, iron corrosion in subsurface PRB systems causes an increase in pH and can generate free Fe2+ and H2

(Equations 4 and 7). Monitoring for increases in pH, the appearance of ferrous iron and increases in dissolved

hydrogen can confirm that corrosion reactions are occurring. The redox status of the aqueous environment should

decline significantly as the corrosion process proceeds. Levels to less than -400 mv versus the standard hydrogen

electrode (S.H.E) have been observed in some installations.

Similarly, when chromate reduction by Fe(0) is occurring the key indicator parameters are pH increase and ferrous

iron formation, but also include the loss of dissolved Cr from solution accompanied by the formation of a mixed

Fe-Cr (oxy)hydroxide mineral phase solid solution (Equations 11 and 12). As pH increases, a shift occurs in

dissolved carbonate equilibria with decreasing concentrations of carbonic acid and bicarbonate species in solution

and increasing concentrations of carbonate ions. The potential for precipitation of calcite and siderite minerals can

be evaluated by monitoring for changes in alkalinity, ferrous iron and calcium. By knowing which reactions govern

contaminant transformation for a given system of contaminant and reactant, the monitoring program can be tailored

to confirm that the system is performing as designed.

7.3.7 Coring for Precipitate Buildup, Microbial Effects

Laboratory column and field tests of Fe(0) have indicated the formation of iron (oxy)hydroxides, iron sulfide

minerals, calcite and siderite as potential mineral phases which might impact iron reactivity and cause a decline in

permeability within an iron treatment zone. More study is needed to evaluate the occurrence, rate of formation, and

importance of these and other potential precipitates in existing emplaced iron systems. For PRBs comprised of other

reactive materials, even less is known and more research is needed. Techniques that can be used to assess

precipitate formation include scanning electron microscopy with energy dispersive x-ray analysis, auger spectroscopy, x-ray photoelectron spectroscopy, laser raman spectroscopy and conventional wet chemical extractions. Over

time there will be less need for the collection of such core data, as the rate of precipitate formation and buildup

become more predictable based on aqueous and solid phase site geochemistry and kinetics.

Microbial activity may also be important in terms of PRB performance, especially for systems dependent upon

biologically-mediated contaminant remediation. Microbial characterization can utilize epifluorescence microscopy

and scanning electron microscopy for microbial identification and enumeration along with fatty acid methyl ester

(FAME) analyses to characterize the bacterial populations upgradient, downgradient and within the wall. In an iron

PRB, the presence of a large reservoir of iron under conditions of suitable pH and substrate availability may

promote the activity of iron and sulfate reducing bacteria and methanogens. Enhanced activity could influence zerovalent iron reductive dehalogenation reactions through favorable impacts on redox potential, the iron surfaces, or

through direct microbial transformations of the target compounds. However, this activity enhancement may come at

the expense of biofouling of the permeable treatment zone. Analysis of these effects under field conditions is under

way.

7.3.8 Hydrologic Testing for Permeability Alteration

Hydrologic changes should also be closely monitored. Head measurements, tracer tests and in-situ flowmeters can

be used to monitor changes in system permeability and alteration of flow paths over time. Flow heterogeneities of

the natural subsurface system should be evaluated as part of site characterization and serve as the baseline for

comparison to post-barrier installation. Significant changes in upgradient flow patterns or head measurements can

result in hydrologic conditions which deviate from assumed design parameters and perhaps affect contaminant

residence time in the reactive material or cause bypass around the reaction zone. Tracer tests performed in locations

where prior tests have already been performed can be very helpful in this regard.
40

8.0 Field Installations

8.1 Chlorinated Hydrocarbon and Chromium Removal in Field Scale Systems

Six of the full-scale PRBs installed to date for ground-water remediation are described below. These six

installations occurred from December 1994 through October 1996. From October 1997 to January 1998, six more

full-scale installations have occurred. The configurations of the first six systems are shown in Figures 11 to 16, and

construction details for these six are provided in Table 7. Summary information on these and other installations can

be found on the internet at http://www.rtdf.org.

8.1.1 Industrial Site, Sunnyvale, California (January 1995)

In January of 1995, after being approved by the California Regional Water Quality Control Board, an iron PRB

was installed at an industrial site in Sunnyvale, California (Szerdy et al., 1996). This in-situ system replaced an

existing pump-and-treat system which was being maintained at a significant cost. The capital cost for the in-situ

system, including the slurry walls used to direct ground water toward the permeable reactive wall, was $770,000.

The reactive wall was 4 feet wide, 36 feet long and 20 feet deep (Figure 11). Since installation, VOC concentrations

have been reported as non-detect from monitoring wells located within the iron wall.

The original pump-and-treat system at the site has been removed and the property has been restored to full

economic use. The monitoring wells provide easy access to the in-situ system for periodic monitoring compliance.

8.1.2 Industrial Site, Mountainview, California (September 1995)

This continuous PRB was installed as a contingency measure for the remediation of residual cis 1,2-dichloroethene

(cDCE) and vinyl chloride (VC) contamination found below the water table at an industrial facility in Mountainview,

California. The water table at the site was 10 feet below ground surface.

A source area containing VOCs was dewatered and excavated to a depth of about 25 feet. As part of the backfilling

procedure, a 44 foot long and 5 foot high zone of iron was placed at the downgradient base of the excavation

(Figure 12). The iron zone was installed in order to treat any residual VOCs which were not removed during the

excavation phase.

A low permeability HDPE liner was used to direct ground-water flow from the upgradient source area through the

iron treatment zone. The HDPE liner was placed on top of the iron zone, and extended up to the high water table

elevation on the upgradient side of the treatment zone. This ensured that any ground water contacting the HDPE

Figure 11. PRB configuration, Sunnyvale, California.



Permeable

Treatment Wall

Cement-Bentonite

Slurry Wall

Historical Range in

Groundwater Flow Direction

Soil Cement-Bentonite Slurry Wall

Pea Gravel

Groundwater Flow

Building

250'

220'

40'


41

Table 7. Construction Details for Six Full-Scale Systems

S e it

I l ndustrial

facility,

Sunnyvale,

Califormia

I l ndustrial

facility,

Mountainview, California

I l ndustrial

facility,

Belfast,

Northern

Ireland

I l ndustrial

facility,

Coffeyville,

Kansas

U , SCG facility,

Elizabeth City,

North

Carolina

G t overnment

facility,

Lakewood,

Colorado

I n nstallation

Date

J 5 an. 1995 S 5 ept. 199 D 6 ec. 199 J 5 an. 199 J 6 une 199 Oct. 199

C - ontaminant & high

conc.

Design

0.5 mg L-1 VC

1.4 mg L-1

cDCE

0.2 mg L-1 TCE

2 mg L-1 cDCE 300 mg L-1

TCE 400 g L-1 TCE

10 mg L-1

TCE

10 mgL-1

Cr(VI)

700μg L-1 each

TCE & DCE

15 μgL-1

VC

R e eactive

Wall Type

F l unnel & Gate E l xcavate & fil R e eaction Vesse Funnel & Gat

Continuous

Trench

Funnel &

Multiple Gate

F l unne

Material

Soil-Bentonite

Slurry

N s ot Applicable Slurry Wall Soil-Bentonite

Slurry

Not Applicable

Sealable Joint

Sheet Pilings

F l unne

Length 2 e 20 ft + 250 ft N t ot Applicabl 1 t 00 ft + 100 f 4 e 90 ft + 490 f N l ot Applicabl 1040 ft tota

N f o.o

Gates

1 e Not Applicabl 1 Reaction

vessel 1 e N 4 ot Applicabl

R e eactive

Material Fe0 Fe0 Fe0 Fe0 Fe0 Fe0

R e eactive

Zone

Height

1 t 0 ft 5 l f 1 t 6 ft in vesse 1 t 1 f A t pprox. 23 f 10-15 f

R e eactive

Zone

Length

3 t 6 ft 4 A 4 f N t 2 t 0 f 150 f 40 ft each

(4 x 40 = 160)

( = 4 x 40 =

160) 4 t ft 4 l .5 f 1 t 6 ft in vesse 3 t f 2 f

Gates differed,

low = 2 ft

high = 6 ft

T s otal Mass

of Reactant

2 s 20 tons 9 s 0 ton 1 s 5 ton 7 s 0 ton 4 n 50 ton No Informatio

T t reatment

Wall Depth 2 s 0 ft 1 s 5 to 20 ft bg 1 s 8 to 40 ft bg 1 s 7 to 28 ft bg 3 to 26 ft bg

10-15 to

20-25 ft bgs

T l ota

System

Length

5 t 06 ft 4 t 4 f A t pprox. 200 f 1 t 000 f 1 t 50 f 1200 f

S l pecial

Features &

Misc.

HDPE atop Fe

to surface

upgradient

directs H2

O

through Fe

Walls direct

H2

O to vessel

inlet, gravity

flow to outlet

downgradient

Two contam.

treated. Chain

trencher with

immediate Fe

placement

Largest of its

kind. Gates

installed using

sheet pile box.

C t K os $ K 770 $ K 100 $ K 375 $ K 400 $ K 500 $1000
42

Figure 12. PRB configuration, Mountainview, California.

 Iron

68 ft

4.5 ft

44 ft

15 ft 15 ft

9 ft

Monitoring Well

New Extraction Well

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yyy yyy

yyy

yyy

yyy

Section

Plan

Groundwater 

Flow Direction

Area of Residual Soil 

Excavation Below Water Table

Gravel

Clean Native Soil

Iron

Clay

Finer

Grain

Soil

HDPE

Liner

(20 mil)

Filter

Fabric

ground surface

5 feet bgs

10 feet bgs

(low water table)

25 feet bgs

5 ft

4 ft

Groundwater

Flow
43

would be forced down and through the iron treatment zone. The entire excavation, including the iron treatment

zone, was covered with clean backfill material.

The estimated construction cost of the iron treatment zone was $100,000 including $60,000 for 90 tons of iron.

The 4.5 foot thick iron zone was designed to treat up to 2 ppm of cDCE and 100 ppb VC.

Monitoring at the site has been difficult because fluctuating water levels and pumpage in the area have resulted in

uncertain flow patterns. It is clear, however, that VOC levels in the iron are significantly below those in the

upgradient source area.

8.1.3 Industrial Site, Belfast, Northern Ireland (December 1995)

A circular in-situ reaction vessel was installed to a depth of about 40 feet at an industrial facility in Belfast to treat

up to 300 ppm of trichloroethene (TCE) and related breakdown products. Two 100 foot long slurry walls direct

water to the inlet of the steel reaction vessel, which is 4 feet in diameter and contains a 16 foot vertical thickness of

iron (Figure 13). Ground water flows by gravity through the iron zone and discharges via a piped outlet on the

downgradient side of the slurry wall. The vessel is equipped with a manhole to access the top of the iron zone,

should periodic scarification of the iron surface prove necessary. The system was designed to provide about 5 days

of residence time.

The total cost of the system, including slurry walls, granular iron, reaction vessel, and engineering was about

$375,000 U.S.

The system was designed to meet ground-water quality criteria of 500 ppb for TCE. These criteria apply to ground

water beneath industrial land slated for redevelopment. Flow rates through the reactor have varied substantially

since its installation, but the TCE levels in the system have decreased to 10s of ppb in the effluent sample ports.

8.1.4 Industrial Site, Kansas (January 1996)

This treatment system was installed at the property boundary of an industrial site in Kansas in order to treat ground

water containing up to 400 ppb of trichloroethene (TCE) and 100 ppb of 1,1,1-trichloroethane (TCA).

The system uses a funnel-and-gate configuration to direct ground water through a single, 20 foot long, 3 foot thick

permeable treatment gate (Focht et al., 1996). The funnel section of the system consists of two 490 feet long soilbentonite slurry walls on either side of the treatment gate (Figure 14). A low ground-water flow velocity of

0.2 ft/day permitted the use of this relatively high funnel-to-gate ratio. The system is installed to a depth of 30 feet

in a basal alluvial aquifer. The treatment gate contains 70 tons of granular iron. The installation cost for the system,

including slurry walls, treatment gate and granular iron, was approximately $400,000.

No determinations of ground-water velocity through the system have been made to date. Concentrations in the iron

zone are below MCLs.

8.1.5 USCG Facility, Elizabeth City, North Carolina (June 1996)

A full-scale demonstration of a permeable reactive barrier (Figure 15) to remediate ground water contaminated

with both chromate and chlorinated organic compounds was initiated at the USCG site by researchers from the U.S.

EPA National Risk Management Research Laboratory (NRMRL) and the University of Waterloo in 1995. A

continuous PRB composed of zero-valent iron was installed in June, 1996 using a trencher that was capable of

installing the granular iron to a depth of 24 feet. The continuous trenching equipment used for the installation was

similar to a large “Ditch Witch” (Figure 9).

This was the first application of the continuous PRB configuration in a full-scale system to treat a combined

inorganic and organic plume. It was designed to meet cleanup goal concentrations of 0.05 mg/L Cr(VI), 5 μg/L

TCE, 70 μg/L DCE and 2 μg/L VC. The trenched PRB was approximately 2 feet thick, 24 feet deep, and about

150 feet long. The PRB begins about 3 feet below the ground surface and consists of about 450 tons of granular

iron. The total installation cost was $500,000, with the cost of iron representing approximately 35% of the total. It

was installed with the trencher in less than 8 hours.

In addition to the compliance wells, the PRB is monitored using a series of multilevel sampling devices to monitor

the geochemical mechanisms occurring in the barrier, and in the downgradient aquifer. To date there have been

2 years of post-installation performance monitoring performed. For all but one quarterly sampling event, 15 multilevel samplers (7 to 11 sample ports per sampler) and 9 to 10 compliance (2 in. 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, cDCE, vinyl chloride,

ethane, ethene, methane, major anions, and metals. In addition, numerous vertical and angle cores have been
44

Figure 13. PRB vessel configuration, Belfast, Northern Ireland.







 Paved

Parking

Lot

100ft 

85ft

Road

GW Flow

Outlet

Slurry

Sand Reactive

Vessel

ground

Inlet

Natural

Geology

40 ft bgs

Granular

Iron

Slurry

Wall

Figure 14. PRB configuration, Kansas.

500ft

500ft

20ft pea gravel



iron zone 

thickness = 3ft

suspected

source area

ground water

flow direction

slurry wall
45

collected to examine changes to the iron surface over time and to evaluate the formation of secondary precipitates

which may affect PRB performance over time. Coring was done vertically (perpendicular to ground surface) and on

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

core through the PRB with the aquifer-iron interfaces intact (front and back of the PRB). These cores continue to be

under study. Inorganic carbon contents, in the form of carbonate minerals, increase dramatically at the upgradient

aquifer sediment-iron interface and decrease within the PRB, reaching background levels within 4 inches

downgradient from the upgradient iron-aquifer sediment interface. Total inorganic carbon content has increased

over time within the PRB.

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

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

Downgradient of the PRB (5 ft), 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 downgradient within the first few meters from the PRB.

Water levels indicate little difference (< 0.3 ft) between wells completed and screened at similar depths upgradient

and downgradient of the PRB, indicating that it continues to effectively function as a “permeable” reactive barrier.

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

6 inches of the PRB as expected. No chromate is detected downgradient of the PRB either in the multi layer

samplers or in the 2 inches compliance wells located immediately behind the PRB.

The vast majority of the multilayer 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 solvent concentrations are highest.

Figure 15. PRB configuration, USCG facility, Elizabeth City, North Carolina.

10 5 0 10

Pasquotank River

pilot 

test area

continous

iron zone

grass border

concrete

Parking area

suspected source area

(plating shop)

flow 

direction

So

150 ft
46

8.1.6 FHA Facility, Lakewood, Colorado (October 1996)

The first funnel-and multiple-gate system using granular iron was installed to a depth of about 25 feet below

ground level at a site in Lakewood, Colorado. The system comprises about 1,040 feet of funnel section (sealable

joint sheet piling) and four reactive gate sections, each 40 feet wide (Figure 16). The gates were constructed using

a sheet pile “box.” Native material was excavated from the box and the reactive material installed, separated from

the aquifer materials by a layer of pea gravel. Capital cost of the system (iron plus construction) was about

$1,000,000. A high degree of lateral geologic heterogeneity and variation in VOC concentrations exists in the

unconsolidated sediments which led to varying iron thicknesses being placed in each gate. Ground water velocities

through the gates were expected to range from 1 ft/day to 10 ft/day, depending upon the hydrogeologic conditions

in the vicinity of the respective gates. Measurements in the PRB gates using a heat-pulse flowmeter have ranged

from < 0.1 ft/day to about 1.5 ft/day (McMahon, 1997).

Design concentrations include up to 290 ppb of trichloroethane (TCA), 700 ppb of trichloroethene (TCE), 700 ppb

of cis 1,2-dichloroethene (cDCE), and up to 15 ppb of vinyl chloride (VC). Half-lives of about 1 hour or less were

measured for these compounds in bench-scale design studies. The only VOC exiting the gates above the 5 ppb

reporting level is 1,1-dichloroethane, which has been measured at up to 8 ppb on the downgradient side of the gates.

There is some evidence of the precipitation of calcite and siderite in the gates based on decreases in calcium and

inorganic carbon in the treated ground water. This is estimated to result in a potential porosity loss of 0.5 percent of

the porosity per year of operation.

Figure 16. PRB configuration, FHA facility, Denver, Colorado.

0 200 400 FEET

Reservoir

Kipling Avenue

GATE 4

GSA-36

GSA-31

GSA-26

GSA-21

GATE 3

GATE 2

GATE 1

flow 

direction

each gate section 40ft in length

sheet pile funnel

gate section

monitoring well data used in design
47

Hydraulic head has increased upgradient of the funnel-and-gate system, with up to 10 feet of head difference

measured across the barrier. This increases the possibility for contaminated water to move around the barrier.

Indeed, VOC concentrations are increasing in ground waters moving around the south end of the barrier and there

is some evidence of VOC moving under the barrier in one location.

8.2 Interpretation of VOC Monitoring Data from the Field-scale Systems

Although VOC concentrations from the six full-scale systems indicate compliance with regulatory criteria, the

monitoring networks installed at most of these facilities do not generate sufficient field data to permit accurate

evaluation of field VOC degradation rates. Pilot-scale systems at Lowry AFB and NAS Moffett Field (installed

under the auspices of AFCEE and the U.S. Navy, respectively), at CFB Borden in Ontario, and at a private facility

in New York state are much better suited to this purpose. Monitoring wells at these sites are located at various

distances in the PRB, as well as on the upgradient and downgradient sides (Figure 17). This allows multiple point

VOC concentrations vs. distance profiles to be obtained, which can be used to calculate field degradation rates.

However, there is significant uncertainty involved in these calculated rates as a consequence of uncertainty in the

ground-water velocity through these systems. Three methods have been attempted to measure ground-water

velocity:

• calculation using water level elevation data and Darcy’s equation

• use of a conservative tracer

• use of an in-well heat pulse flow meter

Tracer tests, with bromide as a conservative tracer, may be very useful but the tests are time consuming.

Calculations using water table elevations are limited by the accuracy of measurement (small gradients over short

distances across the PRB) and the uncertainty in hydraulic parameters (porosity and hydraulic conductivity). The

heat-pulse velocity meter has given magnitudes that are in the ranges anticipated but the directional vectors were,

in some cases, suspect. Because of the ease of use, in-well velocity probes such as the K-V meter, or other in-situ

probes (Ballard, 1996), appear to offer the greatest promise for velocity determinations in PRBs.

A second uncertainty in determining field reaction rates is a consequence of the rapid disappearance of the VOCs

at several sites. This leads to the detection limit being used as the reported concentration at the first sample point in

the reactive zone, with subsequent calculation of a degradation rate from a two point curve (i.e., the influent

concentration and the detection limit). The use of detection limits in a two point curve causes an artificially long

half-life to be calculated since, in reality, the detection limit would be reached at some unknown distance upgradient

of the sampling point.

Degradation rate data from three pilot-scale PRBs are shown in Table 8. These data, when compared to the halflives predicted from the results of bench-scale studies, compare reasonably well given the above uncertainties.

8.3 Inorganic Constituent Removal in Field-scale Systems

8.3.1 Nickel Rim Mine Site, Sudbury, Ontario (August 1995)

A full-scale continuous PRB was installed in an aquifer downgradient from an inactive mine tailings impoundment at the Nickel Rim mine site, Sudbury, Ontario, in August 1995. The PRB was installed by a cut-and-fill

installation technique during which the reactive material was installed within a valley confined by bedrock. The

PRB dimensions are 50 feet long, 14 feet deep and 12 feet wide. The PRB is composed of a reactive mixture

containing municipal compost, leaf compost, and wood chips to promote bacterial sulfate reduction and metal

sulfide precipitation reactions. These organic materials were mixed with pea gravel to attain a permeability greater

than that of the aquifer. Three-foot wide buffer zones containing coarse sand were installed on the upgradient and

downgradient sides of the reactive material.

After passing through the PRB, water quality shows a significant improvement (Benner et al., 1997). Concentrations of sulfate decrease from 2400 - 3800 mg/L to 110 - 1900 mg/L. Concentrations of Fe decrease from 740 - 1000

mg/L to < 1 - 91 mg/L. Alkalinity values increase from 60 - 220 mg/L as CaCO3

 to 850 - 2700 mg/L as CaCO3. The

acid producing potential of the water entering the wall is converted to an acid consuming potential. Concentrations

of dissolved Ni up to 10 mg/L upgradient of the PRB are decreased to < 0.1 mg/L within and downgradient of the

wall. Enumeration of sulfate-reducing bacteria indicates an abundance of these species within the wall, and elevated

numbers in the downgradient aquifer (Benner et al., 1997).
48

Figure 17. Pilot scale PRB configuration, Colorado and New York.

   









Pea Gravel

Iron Filings

Bentonite

Sheet Piling

 Monitoring Well





Pea Gravel

Iron Filings

Bentonite Sheet Piling

 Monitoring Well

Direction of GW flow

Direction of GW flow

Estimated Capture Zone (6.1m)

Estimated Capture Zone (7.3m)



 

15ft

5ft

15ft 15ft

10ft

New York

Colorado

D4 D5 D6

N3 N7

N9

N2

N1

P1 P3

D1 D2 D3

FE1 FE2 FE3

U1 U2 U3

N4 N5 N6

Creek

10ft

9ft

15ft

15ft
49

Table 8. Field Degradation Rates from Pilot-Scale Systems

(A) In-Situ Installation, New York (May 1995)

VOC Predicted Half-Lifea

t1/2, hr

Observed Half-Lifeb

t1/2, hr

TCE 0.4 to 1.1 < 4.0

cDCE 1.5 to 4.0 3.0 to 5.0

VC 2.0 to 6.0 5.0 to 10.0

(B) Pilot Installation Moffett Field

VOC Predicted Half-Lifea,c

t1/2, hr

Observed Half-Lifeb,d

t1/2, hr

PCE 0.6 to 1.2 < 2.5

TCE 1.2 to 1.8 < 0.84

cDCE 6.1 to 9.3 5.7

(C) Pilot-Scale, Lowry AFB, CO (December 1995)

VOC Predicted Half-Lifea

t1/2, hr

Observed Half-Lifea

t1/2, hr

TCE 0.9 to 1.3 2.1 to 4.5

cDCE 4.4 to 6.6 2.6 to 9.3

a - rates at two velocities, temperature adjusted

b - two point curves using detection limit as second point

c - 50% iron

d - 100% iron

8.3.2 Langton, Ontario On Site Wastewater Treatment (July 1993)

A funnel-and-gate system designed to remove PO4

3- and NO3-

 derived from a large-capacity septic system tile field

was installed at a public school in Langton, Ontario, in July 1993 (Charmichael, 1994; Baker et al., 1997). The

funnel consists of two sheet-piling walls extending 12 feet from the central gate area. The gate is 6 feet wide, 5 feet

long and approximately 3 feet deep. It contains two treatment zones, a PO4

3- treatment zone 2 feet thick, and a NO3-

treatment zone 4 feet thick. The PO4

3- zone contains a reactive mixture composed of 6% Fe/Ca oxide material, 9%

high-Ca limestone, and 85% local aquifer sand. Phosphate is removed by adsorption onto Fe oxides and

precipitation of Ca-PO4

 phases. The NO3

-

 treatment zone contains organic carbon in the form of wood chips. Nitrate

is removed by bacterial denitrification.

Monitoring of the performance of the barrier system for over two years indicates that influent PO4

3- concentrations

vary between 1.0 and 1.3 mg/L as P. Effluent concentrations within the treatment gate remained < 0.01 mg/L P for

the first 221 days, then increased to a steady concentration of 0.19 ± 0.04 mg/L P (Baker et al., 1997). Nitrate
50

concentrations varied from 23 to 82 mg/L as N upgradient of the gate, and remained < 1 mg/L as N within the gate

for a 292 day monitoring period (Charmichael, 1994). The very high organic carbon content of the nitrate treatment

zone resulted in the release of high concentrations of dissolved organic carbon and other constituents from this

portion of the gate. Current research focuses on optimizing the reactive mixtures for both PO4

3- and NO3-

 treatment.

8.4 Biological Effects on Field-scale PRBs

8.4.1 Microbial Effects on Iron PRBs

The predominant concern expressed with respect to biological effects on iron PRBs has been the potential for loss

of permeability due to biofouling. This in turn relates to iron-oxidizing bacterial populations that have been

observed in aquifers at other sites and the related plugging of well screens and other treatment equipment. However,

geochemical conditions and bacterial populations in an in-situ PRB of reactive iron are quite different from those

encountered in normal ground-water pumping and monitoring wells. In a well screen, relatively reduced ground

water containing dissolved iron enters an oxygenated environment in the well bore, creating conditions where ironoxidizing bacteria can cause fouling problems. In an iron PRB, ground water becomes even more reducing as it

moves through the iron. The pH within treatment zones of granular iron is generally close to 10, discouraging high

levels of biological activity.

To date, no sliming or other visual evidence of microbial activity has been observed in over 50 laboratory-scale

treatability studies, using ground waters exhibiting a wide range of inorganic and organic chemistry. Though

encouraging, the applicability of these observations across a range of geochemical and microbiological in-situ

conditions is subject to question.

The effects of microbial activity have also been examined at the field scale. Cores of the reactive wall at the

Borden test site, collected two years after installation, showed no significant population of iron-oxidizing microbes,

and only low numbers of sulfate-reducers (Matheson and Tratnyek, 1993; Matheson, 1994). No evidence of

microbial fouling or decrease in performance was observed at the Borden site over a monitoring period of 5 years.

Phospholipid-fatty acid analysis of ground water from an above-ground test reactor at an industrial facility in

California showed no enhanced microbial populations in the reactive iron relative to background ground-water

samples. An above ground reactor has been operating since October 1994 in New Jersey with no indications of

biofouling.

The most detailed microbial ground-water sampling of an in-situ installation has been completed at a pilot-scale

system in New York state. Data on microbial biomass and composition were collected from wells in the iron zone,

in upgradient and downgradient pea gravel zones, and in upgradient and downgradient aquifer monitoring wells six

months following installation. The background microbial community in the aquifer at this site appeared to be

disrupted during construction, and then re-established itself in the new environment created by the gate. There was

no evidence of significant microbial growth in the upgradient gravel and iron zones; the microbial populations in

the upgradient, iron, and background zones appeared to be of similar size and composition. The microbial biomass

in the downgradient gravel and aquifer zones was approximately 10 times greater than the microbial biomass in the

background, upgradient gravel, and iron zones and was of different microbial composition. The difference in

microbial population size and composition can largely be explained by the changes in the geochemistry created by

the treatment wall on the downgradient side of the iron zone, notably the production of hydrogen gas from the iron

which supports the activity of many obligate anaerobic bacteria such as sulphate reducing bacteria. Although the

biomass increased on the downgradient side of the wall, it remained small relative to the microbial biomass

commonly found in surficial soils and shallow aquifers.

Analysis of cores taken from the New York site two years after installation, and cores taken from the Lowry AFB

pilot 18 months after installation, have confirmed the lack of significant microbial activity in iron PRBs. Plate

counts on core samples showed only small numbers of microbial populations. No biofilms were observed on core

samples examined by scanning electron microscopy.

In summary, assessment of microbial activity to date appears to show little effect on the performance of the

reactive zero-valent iron materials at both the laboratory and field scale.

8.4.2 Microbial Effects in PRBs for Inorganic Constituents

Reactive barriers designed to promote biologically mediated reactions exploit the growth and respiration of

anaerobic bacteria. Denitrification barriers are designed to promote biological denitrification by facultative

anaerobic bacteria, such as Pseudomonas sp. These barriers are designed to provide the nutrients required for

bacterial growth and sustained survival, with the exception of dissolved nitrate, which is provided by the plume of

contaminated ground water. Observations of denitrification systems indicate active and sustained denitrification
51

over prolonged periods with no requirement for nutrient addition or mechanical modification (Robertson and

Cherry, 1995). Although no efforts have been made to quantify the bacterial populations of these systems,

continued denitrification indicates that conditions suitable for bacterial growth are sustained.

Barriers designed to promote sulfate reduction and metal precipitation as sparingly soluble precipitates require the

growth and activity of sulfate reducing bacteria. Enumeration of sulfate reducing bacteria in the full-scale reactive

barrier at the Nickel Rim site was conducted annually for the two years following barrier installation. The results of

these studies indicate an abrupt increase in these bacteria within the barrier, increasing to more than five orders of

magnitude greater than in the upgradient aquifer (S.G. Benner, personal communication). Because of changes in

enumeration methods, bacterial numbers derived from the two sampling sessions are not directly comparable.

These two enumerations, however, do show similar trends. In addition to sulfate reducing bacteria, the performance

of the reactive barrier at the Nickel Rim site is affected by the activity of iron reducing bacteria and other

heterotrophic bacteria. Enumerations of these bacteria, and estimates of bacterial activity are under way. In addition

to direct enumeration of bacterial populations and measurements of bacterial activity, the activity of sulfate

reducing bacteria in this system is indicated by measurements of the isotopic ratio of dissolved sulfate, the isolation

of hydrogen sulfide and the identification of iron sulfide precipitates. All of these parameters indicate the

occurrence and activity of sulfate reducing bacteria within the barrier.

8.5 Effects of Mineral Precipitation on Field-scale PRB Performance

Coring activity at pilot-scale PRBs in Canada, New York and Colorado have confirmed the formation of carbonate

precipitates in the upgradient portion of these systems. Aragonite, calcite, iron oxyhydroxides and magnetite have

been identified in core samples. Porosity losses in the range of 10% over the first foot of iron were measured at the

Colorado and New York sites, after 18 months and two years of operation, respectively. Given that the original

porosity of the media was on the order of 0.5, it is not expected that flow patterns have been significantly affected.

The velocity measured at the New York site after two years (immediately prior to coring), was similar to that

measured after six months.

To date, there is no evidence that precipitate formation in PRBs has adversely affected system performance. There

was no discernible change in the performance of the CFB Borden trial over the 5-year period of monitoring. Cores

taken after two years showed no measurable precipitate accumulation and only slight oxidation of the upgradient

iron/aquifer interface was observed after four years (O’Hannesin and Gillham, 1998). Ground-water velocity and

VOC removal rates were very similar at six months and 25 months at the New York installation. VOC removal rates

appeared unchanged after 6 months and 18 months of operation of the pilot-scale PRB at Lowry AFB in Colorado.

Removal of precipitates could represent a significant operations and maintenance (O&M) cost for PRB technology. Though the evidence indicates that precipitates do not have a significantly adverse effect on reaction rates, they

nevertheless form and one must assume that they will eventually cause an adverse reduction in permeability. Based

on numerous laboratory evaluations of porosity loss, and fewer evaluations at field sites, it is estimated that in

highly mineralized and/or oxygenated ground water some degree of maintenance could be required as frequently as

every five years in order to manage potential problems caused by precipitate formation. In less mineralized waters,

the frequency could be as low as every 10 to 15 years. The certainty in these estimates will increase as the period

of experience grows.

To date, there has been no need for rejuvenation, thus methods have not been developed or tested. Methods that

include techniques such as hydraulic or mechanical mixing are being considered. Hydraulic mixing with, for

example, jetting equipment would provide the potential for adding chemical descalants. Though the need for

maintenance to control the effects of precipitate formation remains uncertain, as does the most cost-effective

methods for such maintenance, periodic O&M requirements should be included in long-term cost models for the

technology.
52
A - 53

Appendix A

Summary of Published Results Relevant to Subsurface

Remediation Using Permeable Reactive Barriers
ry of published results relevant to subsurface remediation using permeable reactive barriers

ant Type test Description/Conditions Results Reference

Batch,

Column

US is co-applied in batch/column to remove deposits from metal surface. Column 80:20 sand/iron with 50 mesh Fe,in N2. Various concentrations of TCE used in studies.Steady flow conditions maintained throughout thecolumns with minimal channeling. HPLC pump used tosample from columns to provide accuracy at low flowrates and ensure minimal evaporation.

Batch: US removed inactive deposits and impurities thusextending activity of the metal surface. US also found to cleaveCl-H bond in TCE giving an added benefit . US probe in columndid not extend deep enough but modifications in design showpotential for application. Results w/o US in column: TCE t1/2 =260 min at 3mL/min; no less than 20 PV in the life of thecolumn. t1/2 = 360 min at 2 mL/min; samples taken after > 150PV.

Afiouni, G.F., et al., 212thNational ACS Meeting,Orlando, FL, 36:22-29 (1996)

ene Batch Batch experiments to investigate nitro reduction by

granular iron in model systems. Nitrobenzene (disappears in few h) 

→ nitrosobenzene 

→

aniline Reduction by surface; dissolved Fe2+ & H+ producedduring corrosion.

Agrawal, A. & P.G. Tratnyek,207th ACS Nat’l Meeting, SanDiego, CA, pp. 492 (1994)

ene

nate

Batch Adsorbed H2CO3 & HCO3

- drives metal dissolution by:

Fe0 + 2H2CO3(ads) → Fe2+ + 2HCO3

-

(ads) + H2(g) 

Fe0 + 2HCO3

-

(ads) 

→ Fe2+ + 2CO32-(ads) + H2(g)

Corrosion rates decrease from carbonate precipitates.

Anaerobic bicarbonate buffer, Fluka Fe, and nitrobenzene asoxidant in batch exp. Kobs declined with increased carbonate andextended exposure of metal to carbonate buffer. FeCO3aggregates formed on metal surfaces when using bicarbonatemedium but were not observed on surfaces exposed to DI water.

Agrawal, A., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:720 (1995)

ene Batch 18-20 mesh Fluka Fe turnings, sonicated in 10% HCl,

washed with buffer to remove acidity or Cl-. Anaerobicbatch in 60 mL serum bottles with 2 g dry, sieved Fe.

Rates for nitrobenzene (0.035) 

→ nitrosobenzene (0.034) 

→

aniline 0.008 /min. Rates controlled by mass transfer to metalsurface. Precipitation of siderite on metal surface inhibited nitroreduction.

 Agrawal, A., P.G. Tratnyek.

ES&T, 30(1):153 (1996)

DCE Batch 10 g Fisher pretreated 40-mesh filings; SA 1.0 m2/g. 0.20

g powdered pyrite (buffer), DI water. cDCE & tDCE attwo Ci in ZHEs, anaerobic, at 12 rpm, 22-25°C. 

Reductive dechlorination & sorption; Cl- 80-85%; products:ethene, ethane; more VC found in cDCE. cDCE not 1st orderreaction. Sorption is described by Freundlich isotherms (tDCEsorption > more soluble cDCE); quasi-equil. 1.1 h.

Allen-King, R.M., et al.,Environ. Toxicol. Chem.,16:424-429 (1997)

Batch Aging & concentration effects of acetylene & CCl4 in

Ar-purged vials w/ 100 mL DI H2O, 5 g HCl pretreatedFisher Fe & 0.10 g powder pyrite (buffer), 20EC, 6 rpm. Fresh systems: Fe/H2O not mixed prior to compoundexposure.

 CCl4 : initial rates in fresh systems were 2 to 4 X > in agedsystems and faster at lower concentration. Acetylene initial rate(0.1 to 2.0 

Fmol) were > order of magnitude lower in systems

aged as little as 1 d. For both compounds, pseudo-first-order rateconstant independent of concentration in sufficiently agedsystems (3 to 7 d). Fast reacting sites eliminated within few h dueto precipitation, sorption, corrosion.

Allen-King, R.M., et al., 213thNational ACS Meeting, SanFrancisco, CA, 37:147-149(1997)

te Batch,

Column

Batch and column with Hanford sediments to predict (1)longevity of dithionite, (2) efficiency as reductant of Fe(3) longevity and reactivity of the reduced Fe

Other than initial reaction with ferric iron, primary factorpromoting loss of dithionite in system was disproportionation viaheterogeneous catalysis at mineral surfaces.

Amonette, J., et al., In SituRemed.: Sci. Basis for Current& Future Technol. Sym.Battelle Press, pp. 851 (1994)
ed...

ant Type test Description/Conditions Results Reference

Batch Batch experiments using sand collected from a shallow

sand and gravel aquifer to determine reduction of Cr(VI).Ci Cr(VI) varied 2 to 200 FM; Adjusted pH 4.5, 5.5, 6.5.Note: Fine fraction Fe(II) minerals (#64 Fm dia.)dominated Cr(VI) reduction due to greater reactivity andSA. However, to provide consistent results fines and sandshad to be separated and parallel experiments run on each.

As pH decreased (6.4 to 4.5), Cr(VI) reduction increased 30 to 50nmol/m2 for the sand fraction (645 - 1000 Fm) and 130 to 200nmol/m2 in fines. Amount reduced in both sand and finesincreased from 35 to 80 and from 130 to 1000, respectively, for a10-fold increase. Consistent rate descriptions achieve byassuming that intraparticle diffusion limited the observed rate ofreduction.

Anderson, L.D., et al., ES&T,28(1):178 (1994)

s In situ-Otis

AFB, MA

Induce fracture to fill w/ Fe filings w/ slurry mixturemoving down and outward, creating series of overlappingvertical planes thus becoming a "continuous" wall.

Plan 2 parallel 50’ walls, 2’ apart, perpendicular to flow path of600’ wide plume with 5 to 150 ppb TCE & PCE. Walls willbegin 80’ bgs to top of plume, extending to as deep as 150’.

Appleton, E.L., ES&T,30(12):536A (1996)

nated

s

Batch Al coupon size 2 x 15 x 0.2 cm, Zn & Fe coupon 2.5 x 15

x 0.1 cm. 65 mL solvent and 5 mL DI water in 125-mLflask containing solvent and metal coupon.

Reactivity accelerated when water added. Problems with Al andZn, but not Fe corrosion in dry systems. 1,1,2-trichloroethaneonly structure with appreciable oxidative breakdown.

Archer, W. & E. Simpson,Industrial Eng. Chem. ProductR&D, 16(2):158 (1977)

Batch 1 mL CCl4 heated at 200EC, sealed tubes, for 14 d using

twice the metal shot, powder, granules, or chips needed tocomplete reaction.

Hexachloroethane: end products perchloroethylene, hexachlorobutadiene. Most reactive Al (100%), Ti (100%), Cd (74%),antimony (58%), In (58%), Ge (47%), Sb (33%).

Archer, W.L. & M.K. Harter,Corrosion: Nat'l Assoc. ofCorr. Eng. (NACE), 34(5):159(1978)

CA Batch 5 mL inhibitor-containing solvent refluxed with 0.5 g of 

16-32 mesh pure Al pellets in open reaction tubes. Tubesin oil bath at 7EC. Upper portion extends through a watercooled Al block that acts as a condenser.

Inhibitors compete with solvent for AlCl3 produced at microcorrosion sites by complexing the chemisorbed AlCl3 product.Resultant complex insoluble in solvent, acting as a plug coveringoriginal reaction site.

Archer, W.L., Industrial Eng.Chem. Product R&D, 21:670(1982)

rates;

s

Batch Deoxygenated water (buffered), Zn(0), PCE or TCE.

Sampled for product formation. Early heterogeneousprocess, but initial rate does not increase linearly withincreasing Ci (expected for pseudo-1st-order system), butlevels off as Ci increased.

Reductive elimination (RE) important in Zn(0). ~ 15% PCE →dichloroacetylene (0.25 → acetylene bypass VC). TCE →acetylene (20% of original TCE); trace VC. Chloroacetyleneintermed. → acetylene preferred over VC. Manipulating REover initial hydrogenolysis would be beneficial goal.

Arnold, W.A. & A.L. Roberts,213th National ACS Meeting.San Francisco, CA, 37:76-77 (1997)

Column Limestone or sand 2 cm over 10-cm depth of soil in PVC

column. Leachate passed through columns at 1 PV/24 huntil breakthrough. Unamended leachate diluted to 25%and pH 4.0 or 2.5 to keep Cr in solution. ~3,000 ppm TOCupon dilution and adjusting pH to 2.5. 

Limestone delayed breakthrough of Cr. Retained Cr(III) morethan Cr(VI). Retention>> Be, Cd, Fe, Ni, Zn. TOC & Fe(II)determine Cr(VI)/Cr(III) (Clay, Fe oxides better at retaining Cr).pH affects solubility of Cr and limestone. When Cr(VI) & Cr(III)in leachate, migration delayed several-fold by limestone barrier. 

Artiole, J. & W.H. Fuller,Journal of EnvironmentalQuality, 8:503-510 (1979)

Batch Zinc powder ± B12 under N2. Initial conc. 2.2 mM CCl4. Zn + B12 dechlorinated CCl4 to CH4 (50% recov). Rates slowed

when B12 absent [CCl4 → CHCl3, DCM, CM, CH4 (80%recov)].

Assaf-Anid, N. & L. Nies,209th Nat’l ACS Meeting,April 2-7, Anaheim, CA,35:09-811 (1995)

ro

HCB) Batch Zinc powder ± B12 under N2. Initial concentration of 50 FM HCB. Product pentachlorobenzene higher rates w/o B12 at 9.6 h-1 compared to 0.3 h-1 w/ B12. B12 may compete w/ HCB for e-.

over initial hydrogenolysis would be beneficial goal.

(1997)

Column Limestone or sand 2 cm over 10-cm depth of soil in PVC

column. Leachate passed through columns at 1 PV/24 huntil breakthrough. Unamended leachate diluted to 25%and pH 4.0 or 2.5 to keep Cr in solution. ~3,000 ppm TOCupon dilution and adjusting pH to 2.5. 

Limestone delayed breakthrough of Cr. Retained Cr(III) morethan Cr(VI). Retention>> Be, Cd, Fe, Ni, Zn. TOC & Fe(II)determine Cr(VI)/Cr(III) (Clay, Fe oxides better at retaining Cr).pH affects solubility of Cr and limestone. When Cr(VI) & Cr(III)in leachate, migration delayed several-fold by limestone barrier. 

Artiole, J. & W.H. Fuller,

Batch Zinc powder ± B12 under N2. Initial conc. 2.2 mM CCl4. Zn + B12 dechlorinated CCl4 to CH4 (50% recov). Rates slowed
ed...

ant Type test Description/Conditions Results Reference

rus Column,

Cylinder, 

Reactive

Wall

Permeable mix 50% sand, 45% crushed limestone, 5%metal oxide in acrylic column w/ 3.3 mg/L PO4 -P over 3y (1250 PV). Biofilter effluent in 0.5 x 0.5 m cylinder 2L/d over 133 d (101 PV). Funnel-&-gate in septic plume.7 m long funnel, 1.8 long; 2 wide; 10 deep (m) gate for779 d.

>90% efficiency in column & cylinder. PO4-P 0 to 0.3 mg/L incolumn effluent. Phosphate accumulated on oxide surfaces &precipitated as microcrystalline hydroxyapatite. Cylinder: 3.93 to0.14 mg/L-P; 2.50 to 0.05 mg/L-ortho P. F&G: 4 m above 2 to 3mg/L; 0.4 m above gate 1.5 to 0.4 mg/L. Average phosphate inwall ~ 0.19 mg P/L.

Baker, M.J., et al., Internat’lContain. Technol. Conf. &Exhib., St. Petersburg, FL, Feb9-12, pp. 697 (1997)

 VC,

o

e

Batch,

Column

Site ground water (GW) from DOE Pinellas Plant. VWRcoarse iron filings used for high reactivity and low cost.

Batch: fast rates for TCE, DCE , VC in site GW. Dichloromethane rates very slow. Column: t1/2 TCE = 36-103 min; DCE =150-200 min. However, rapid plugging of iron by Pinellas GW.

Baghel, S., et al., G. E. Corp.R&D Center for USDOE,Sandia (1995)

CE Containers,

TN & OH

55-gal drums of Na-chabazite zeolite (remove Sr, Cs. atseep, Oak Ridge National Lab, TN) and using 3 mediatypes (reduce TCE, Portsmouth Gas.Diff. Plant, OH).

>99.9% Sr, Cs removal ORNL (25% red. total radioactivedischarge). TCE reduced at PORTS X-120 site. Drums predictflow and allow easy media replacement and monitoring.

Barton, W., et al., Internat’lContain. Technol. Conf. &Exhib., St. Petersburg, FL, Feb9-12, pp. 827 (1997)

ne

ge

Reactive

Wall,

Ontario

Wall installed at the Nickel Rim mine site near Sudbury,Ontario on 8/95. 15 m long perpendicular to GW flow, 3.6m deep, 4 m thick. Used municipal & leaf compost, woodchips, and pea gravel (for permeability). Sand buffer, claycap. Mon. wells parallel to GW flow. 

After 9 mo sulfate reduction and metal sulfide precipitation.Sulfate: 2400 to 4600 mg/L to 200 to 3600 mg/L, Fe:250 to 1300mg/L to 1 and 40 mg/L; pH 5.8 to 7.0. Alkalinity rose 60-220 to850-2700 mg/L as CaCO3. Fe mono-sulfide precipitate. Cost ~$30,000.00 (half materials/half installation) potential life 15 y.

Benner, S.G., et al.,1 ACS Meeting, San Fran.,

CA, April 13-17, 140(1997)

2 Internat’l Contain. Technol.

Conf. & Exhib, St.Petersburg, FL, 835 (1997)

3 GWMR, Fall, 99-107

(1997)

Reac. Wall

Eliz. City,

NC

Models indicate reactive barrier most efficient/costeffective. 46 long x 0.6 wide x 7.3 deep (m) barrierinstalled < 6 h using continuous trencher 6/22/96.

11/9 sampling indicated 2.5 mg/L Cr(VI) declined to < MCLwithin barrier. Further sampling under way determining groundwater chemistry and organic concentrations.

Bennett, T.A., et al., 213thNational ACS Meeting, SanFrancisco, CA, April 13-17,pp. 243-245 (1997)

 PCE

opyl

xtrin

D)

Batch,

Column

Batch: 100 mesh Fisher Fe powder & -8 to 50 mesh Peerless mix. 40 mL sealed vials, 10 g Fe, PCE, 40 mg/LCaCO3; 0, 45, 70 g/L HP-β-CD. Column: 0.47 m x 5 cm,15 cm 70/100 mesh sand below (30-cm head) 10 cm 30/40mesh sand. 25 mL PCE to form pool. RT 25.3 h.

HP-β-CD enhances solubility w/o decreasing interfacial tensionof PCE and water. Smaller % PCE degraded at higher HP-β-CDconc. PCE decreased in both non-recycling and recycling of posttreatment effluent. Greater degradation at higher iron SA. Plan touse higher Fe SA and longer RTs to increase degradation.

Bizzigotti, G.O., et al., ES&T,31:472-478 (1997)

Batch,

Column

Batch: Compare 100 g siderite, pyrite, Fe(0) (~ 0.5-1mm), chips (~ 1-5mm). Use 500 g Cr in CaCO3 DI wateragitated at room temp. Column: 15 x 6.5 cm, pump tracerbase upward to determine void v & dispersivity ofcolumn.

Batch: Fe(0) > pyrite. Column: Fe chips (remove 50% Cr < 2 h) >pyrite (w/ calcite 50% 0.5 h; no calcite 50% 1 h) > coarse Fe(50% 28 h). Chips: Coating (rf 5.2). No Cr 4.5 PV. Filings: No Cr> 15 PV. Remain active, little coating.

Blowes, D.W. & C.J. Ptacek,Subsurface Restor. Conf., 3rdInterna’l Conf. on GroundWater Quality Res., June21-24, Dallas, TX, 214-216(1992)

 etc Trench Excavate trench, fill with active material such as pyrite or

elemental iron to transform & precipitate contaminant.

This patent relates to the treatment of GW for the purpose ofremoving water-borne contaminants.

Blowes, D.W. & C.J. Ptacek,Patent 5,362,394 (1994)

Tile

ge

In-Line

Bioreactor

Two 200-L fixed-bed bioreactors, with coarse sand andorganic carbon (tree bark, wood chips and leaf compost),to treat 3-6 mg L NO3 -N from farm-field drainage tile.

Reduced NO3-N < 0.02 mg/L at 10-60 L/d over a l-y period byanaerobic denitrification promoted organic carbon. Design issimple, economical and maintenance free.

Blowes, D.W., et al., Journalof Contaminant Hydrology15:207-221 (1994)
ed...

ant Type test Description/Conditions Results Reference

Batch,

Column

Batch: 100 g (50% fine Fe filings, 49% sand, 1% calcite),Cr(VI). 2nd mix 50:50 Fe, quartz. Column: 50% Fe(0),50% sand, top 5 cm 1% calcite. 20 mg/L Cr(VI) insimulated GW.

Cr(VI) 25 to <0.05 mg/L 3 h batch. Column: No breakthrough Crafter 140 PV. Dissolved & total Cr < 0.05 mg/L. Fe(III)oxyhydroxides form but not sufficient to inhibit Cr(VI) reductionat experimental velocity.

Blowes, D.W, et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:780 (1995)

ne

ge

Test

Reactive

Wall

Test cell 1.5 long x 1 wide x 1 deep (m) installed 10/93 insand aquifer ~ 75 m downgradient tailings impoundment.Organics (leaf, pine mulch, bark), creek sediment (sulfatered. bacteria), limestone, coarse sand and gravel.

1 m on flow path, SO4 3500 to 7 mg/L, Fe 1000 to < 5 mg/L, pH& alkalinity increased from sulfate-reduction. Sulfate-reducingreactive walls are potentially effective and economical solutionto many acid mine drainage problems.

Blowes, D.W., et al., Mining& Environ. Conf., CANMET.Sudbury, Ontario, May 28 toJune 1, 3:979 (1995)

CE Fe

Reactive

Wall

Wall of 100% Fe filings 46 m long, 0.6 m wide, 7.3 mdeep installed in < 6 h using a continuous trenchingtechnique at Elizabeth City, NC on 6/96. Bench labstudies, and flow and transport models used in walldesign.

Site GW, reactive materials: Cr(VI) 11 to <0.01 mg/L and TCE1700 Fg/L to < 1 Fg/L. Decrease of Cr(VI) from influent 6 mg/Lto < 0.01 mg/L and TCE from 5600 Fg/L to 5.3 Fg/L withinwall. TCE approaches or attains the MCL within the barrier.

Blowes, D.W., et al., Internat’lContain. Tech. Conf. & Exhib.St. Petersburg, FL, Feb 9-12,pp. 851 (1997)

Batch,

Column

Batch: 500 g Cr(VI) to 100 g solid mixes (siderite, pyrite,coarse & fine Fe(0)) in open flasks, agitated, room temp.Settle 5 min; 10 mL sampled. Column: 6-6.5 cm diaacrylic, 5-15 cm long, 1-20 cm long with layers reactivemix. Void volume & dispersivity determined. Cr(VI)solution introduced.

Batch: Rate of Cr(VI): fine Fe(0) > pyrite & coarse Fe(0) .Column: partial reduction of Cr(VI) by pyrite & coarse Fe(0);quantitative reduction of Cr(VI) by fine Fe(0) at rapid velocities.Fe(0) reduces Cr(VI) to Cr(III) with oxidation of Fe(0) to Fe(II)& Fe(III), and precipitation of sparingly soluble (oxy)hydroxide.Cr(III) forms a solid solution or adsorbs on goethite.

Blowes, D.W., et al., ES&T,31:3348-3357 (1997)

CT,

DCM

Batch Open circuit potential time measurements using Fe with

CT & DCM in borate buffer and simulated GW of KBr &CaCO3. Polarization of Fe electrode in borate solution towhich 0.2 mL of CT added. CT acts as an oxidizer of Feelectrode, while DCM does not.

Injection of CT shows faster and larger potential shifts in Fe thanin freshly cleaned Fe electrode. Potential decay in all GWstudies. Magnitude depended on pH and solution. Borate andKBr decay mainly from chemical dissolution of films. In CaCO3,autoreduction/chemical dissolution may be responsible. 

Bonin, P.M.L., et al., 213thNational ACS Meeting, SanFrancisco, CA, 37:86-88(1997)

Batch Vaporization procedure (SMAD or cyro method) to

compare metal powders.

H2O oxidation overwhelmed Mg-CCl4 reaction. Sn, Zn degradeCCl4 but differ in carbon product (Zn → CH4; Sn → CO2.Intermediate Cl3CMCl may be protonated by H2O to give CHCl3or eliminate CCl2 which subsequently reacts with H2O to formCO2 and HCl.

Boronina, T., et al., ES&T,29:1511 (1995)

Column Simulated process & GW from DOE uranium enrichment

plants. Packed column of steel wool, Dowex TM 1-X8.

Use of iron economical but may be more difficult to accuratelypredict its sorptive capacity or functional "lifetime"

Bostick, W., et al., Oak RidgeK-25 Site Rep. MartinMarietta. DOE K/TCD-1141(1995)

Ni,

O22+

Batch Shake solid w/ soln 16-24 h in sealed container. Exp 2:

0.01, 0.03, or 1.0-g iron to 10mL w/ 8 mg/L U, shake 18h, sample day 1 & 30. Exp 3: 1.43 x 0.15-cm iron couponsin 500-mL bottle w/ 927-mg/L U. N2 purge.

Iron surfaces passivate at elevated pH (little activity at >9.5).Sorption to iron corrosion products predominant removal processfor uranyl (Cd, Mn and Ni also). Sorbed products need to becontrolled. Columns needed to determine long-term capacity.

Bostick, W., et al., Oak Ridge& Martin Marietta EnergySystems, Inc. for U.S. DOE. K/TSO-35P (1996)
ued...

inant Type test Description/Conditions Results Reference

Batch Continued from previous reference. Fe(0)w/ sand or pelletized Fe oxide most effective. Reduces

cementation from rust; enhances dilution of hydroxyl ionreaction product; enhances sorption of cationic contaminant topelletized Fe.

Bostick, W., et al., Oak Ridge& Martin Marietta EnergySystems, Inc. for U.S. DOE. K/TSO-35P (1996)

Batch Batch: ~1.4 x 0.16 cm Fe coupons, simulated GW, soluble

U as uranyl nitrate. Gas glove box. Pure O2 added to yieldequivalent to solution purged with lab air.

Under oxic conditions, U(VI) rapidly and strongly sorbed tohydrous ferric oxide particulate (?rust"), whereas U slowly andincompletely reduced to U(IV) under anoxic conditions.

Bostick, W.D., et al., Internat’lContain. Technol. Conf. &Exhib, St. Petersburg, FL, Feb9-12, pp. 767 (1997)

PCE Batch 4-15 mL vials, anaerobic: 2 controls/2 Fe+pyrite (5 g pretreated Fe filings, 0.1 g ground pyrite [buffer]). Reaction orders 2.7 TCE & 1.3 PCE total system conc. Nonlinear sorption fit generalized Langmuir isotherm. 1st order

rates.

Burris, D.R., et al., ES&T,29:2850 (1995)

E Batch 15 mL serum vials, 5 g iron, 0.1 g pyrite (ground). Pyrite

added to stabilize the pH to 6.5 - 7.0. Anaerobic

Rapid initial rate followed by a slower rate. Sorption of PCE toFe(0) follows Langmuir-type isotherm. 

Campbell, T.J. & D.R. Burris.209th Nat’l ACS Meeting,Anaheim, CA, 35:775 (1995)

MoO42-,

CrO42-

Batch Kinetic studies in 50 mL redox-sensitive-metal solutions

added to polystyrene centrifuge tubes containing 1-g 40mesh metallic iron with SA of 2.43 m2/g.

Particulate Fe(0) effectively removed each of the contaminantsfrom solution by reductive precipitation. Removal ratesdecreased by CrO42- > TcO4- > UO22+ >> MoO42-.

Cantrell, K.J., et al., J. of Haz.Mat., 42:201 (1995)

ous Batch 0.2% Fe colloids, surfactant. Polymers (vinyl (VP), vio

(GX), cellulose (CMC )) tested to increased colloidalFe(0) mobility in porous media. Turbidimeter meas. Fecolloids.

VP is superior to GX and CMC because VP suspension producedthe lowest back pressure, resulting in the highest hydraulicconductivities.

Cantrell, K.J., et al., J . ofEnviron. Eng.-ASCE, 23:786(1997)

ous Column 1 m columns, 4.4 cm dia. 20-30 mesh sand avg 0 = 0.32.

Fe(0) colloid dia. = 2 Fm, bulk density = 2.25 g•cm-3,particle density = 7.6 g•cm-3. V = 0.154, 0.307, 0.614cm/s. 0.01M CaCl2 at 0.2 cm/min (~2 PV) to simulateGW.

Colloidal-size Fe(0) injected into porous media forming chemicalreactive barriers. Relatively even distributions of Fe(0) in sandcolumn at low conc.; high injection rates. As V increased,distribution of Fe(0) colloids became increasingly even.

Cantrell, K.J. & D.I. Kaplan, J.of Environ.Engineering-ASCE123:499-505 (1997)

phenol Batch Palladized graphite and carbon cloth electrode in 3-neck

round-bottom flask (50 mL). Cu wire threaded through 5mm glass tubing. Central portal used for pH & cathodepotential. Anodic compartment vented for O2 escape.

Rapid dechlorination of 4-chlorophenol on palladized carboncloth or palladized graphite electrodes. Hypothesize H2 gasinterlaced in Pd lattice powerful reducing agent dechlorinatingcompounds adsorbed on palladized electrode surfaces.

Cheng, I.F., et al., ES&T,31:1074-1078 (1997)

ate Batch 4 g untreated 325 mesh iron to 50 mL of 12.5 mM nitrate

buffered at pH 5. No effort made to exclude O2.Vigorously stirred.

Nitrate < 0.2 mM in 74 min. Ammonia 103%. Pseudo-1st orderrate constant 0.0530/min. Buffer is key in nitrate reduction.

Cheng, I.F., et al., 213th ACS,San Fran., CA, 37:165 (1997)

Aroclor

1254)

Batch Iron powder (0.5 g), 1.5 Fmol PCBs in 10 mL flamesealed glass ampule. PCBs undergo dechlor. & other reactions at > 300EC in presence of Fe powder. Virtually complete loss of chlorinated congeners. Chuang, F.-W. & R.A. Larson, ES&T, 29:2460 (1995)
ed...

ant Type test Description/Conditions Results Reference

 PCE Batch Solutions incubated with lab grade iron under static

conditions. Concentrations monitored as a function oftime. Pyrite with Fe to counterbalance pH increases. 

Oxidation reactions of Fe increased pH. Halocarbons convertslower at high pH. High grade pyrite more reactive. SA importantconsideration.

Cipollone, M.G., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:812 (1995)

Column Long-term (~30 d, >300 PV) & fast velocities (5-30

cm/h) w/ 4-600 mL Fe mix columns. Rates varied bymedium, SA, mixture, time. Influent 50 FM TCE at 3.15mL/min.

Fe oxidation increases pH, pyrite decreases pH. Pyrite is placedat column head because precipitation slower at lower pH w/wider dispersion of precipitation zone throughout column andlessens plugging at head of column.

Cipollone, M.G., et al., 213thACS Nat’l Meeting, San Fran.,CA, 37:151-152 (1997)

Batch Fe(II), in slightly acid to base solution, in 500-cm3 bottle

w/ hydrophobic inner surface, ambient temperature, andunder anaerobic conditions. 0.1 FeCl2 (pH 4) added,sampled by syringe through septum.

3 e- reduction process, although thermodynamically feasible,slow if at all. Fe(II) [sorbed to wall or precipitate as Fe(OH)2(s)or FeCO3(s)] reduced TcO4-. Rates proportional to sorbed orprecipitated Fe(II). Discuss redox/paths TcO4- to TcO2•nH2O.

Cui, D. & T.E. Eriksen,ES&T, 30:2263-2269 (1996)

und

bon

on

Batch 15.0-mL sealed vial. 5.00 g Fe with low CO2 water,

anaerobic, 8 rpm in dark at 20EC. pH increased 5-7 whenFe added. 13CO2 to determine if CO2 is being reduced byiron to form the hydrocarbons. Note: Non-Clhydrocarbons form during ethene reduction and whenthere is no ethene. Fischer-Tropsch type synthesis ofhydrocarbons, proposed for hydrocarbon production.

13CO2 was not incorporated into hydrocarbons produced. Aciddissolution of gray cast irons containing both carbide andgraphite carbon yielded hydrocarbons and a substantial amountof graphite residual. The dissolution of metallic irons containingonly carbide carbon yielded total carbon conversion tohydrocarbons. Carbide carbon in the iron most likely carbonsource for the production of the background hydrocarbons.

Deng, B. & A.T. Stone,ES&T, 30:463-472 (1996)

Batch 40 mesh HCl pretreated Fisher Fe filings, SA 1.18 m2/g

(0.2 to 10 g) using 15 mL borosilicate in ZHE filled byVC solution, anaerobic, 8 rpm, 20EC. Also at 4, 20, 32,45EC.

5.0 g Fe/15.0 mL: VC → ethylene (Partial absorb. to Fe). Ratesincrease as Fe & temperature increases. Activ. E 40 KJ/moleindicate surface reaction. H2 also produced, but, Fe2+ not directlyinvolved in reduction.

Deng, B., et al., 213th NationalACS Meeting, San Francisco,CA. 37:81-83 (1997)

Pilot

barrier in 

CO

Pilot barrier using Fe to remove U from tailings effluent.at UMTRA site scheduled 5/96 to 1999. Fe foam SA 0.1& 5 m2 /g; Fe(0) SA 5.6E-3 m2/g. During 3 years U, Se,Mo, other elements monitored as well as costs andbenefits.

Field coincides with experimental findings. In Fe foam batch Uremoved to <DL in 10 h. Metallic hydraulic conductivitymaintained. Initially conduc. 6.4 x 10-3 cm/s; Fe foam was 0.53cm/s. Capacity maintained after > 700 PV oxygenated water.

Dwyer, B.P. & D.C. Marozas,Internat’l Contain. Tech. Conf.& Exhib. St. Petersburg, FL,Feb 9-12, pp. 844-850 (1997)

D2O Column Glass column with wet Ottawa sand or VWR coarse iron

filings. 2.65 g/cm3 for sand and 6.5-7.6 g/cm3 for iron.

2 tracers (D2O, KBF4) for Fe column. D2O more conservative andKBF4 easier on-line detection. Both inert with respect to Fe.

Eykholt, G., et al., 209th Nat’lACS Meeting, Anaheim, CA,April 2-7, 35:818 (1995)

r,

lor

Batch

Kinetics

100 mL ZHEs, ~ 40 g coarse Fe filings (40 mesh, SA 13.5m2/g), 10 mg/L or 100 mg/L alachlor and/or metalochlor,room temp, 3 rpm. Sampled over even intervals for 5 d.

Rapid dechlorination by Fe(0) shown by Cl- and GC/MSanalyses. Apparent 1st-order kinetics, but indication of ratelimited and instant sorption. 2-site batch kinetic model fitted toresults.

Eykholt, G.R.& D.T.Davenport, 213th Nat’l ACSMeeting, San Francisco, CA,37:79-81 (1997)

Elemental

Fe cathode

Experimental reactors using a two-part glass vessel with aNafion-117 proton permeable membrane.

Reduction of CCl4 →chloroform in hrs at .0005 to .005 / min.Suggest H2 serves as intermediate for CCl4 hydrogenolysis.

Festa, K.D., et al., 209th Nat’lACS Meeting, Anaheim, CA,April 2-7, 35:711-715 (1995)
ed...

ant Type test Description/Conditions Results Reference

loro

e Column 50 & 93 mg/L 1,2,3-TCP was passed through 6 flow- through reactive columns containing Fe(0), silica sand &

simulated GW.

End-product propene. Fe(0)enhanced dechlorination. Rateincreased in proportion to iron SA to solution volume ratio. Ratios of 1.16, 3.7, 8 m2/mL gave t1/2 17.6, 6.6, 3 h, respectively.

Focht, R.M. & R.W. Gillham.209th National ACS Meeting,Anaheim, CA, April 2-7,35:741 (1995)

w Reactive

Walls

Gillham founded EnviroMetal Technologies in 1992 tocommercialize the reactive wall technology.

To Date 5 full-scale in situ treatment zones installed: 2 incommercial sites in CA, 1 in KS, 1 in N Ireland and 1 atElizabeth City, NC

Focht, R.M., et al.,Remediation, Summer:81(1996)

Funneland-Gate In Sunnyvale CA site, 100% pure granular iron filing wall is 4’ thick, 40’ wide and 20’ deep. TCE levels of 30-68 ppb entering wall reduced to < 0.5 ppb; cDCE of 393-1916 ppb to < 0.5 ppbFunneland-Gate 1000 foot funnel- and-gate system installed at industrial facility in Kansas in January 1996 to treat 100 to 400 ppb

TCE. Reactive zone 30' to 17’ bgs and 3’ thick.

Under optimum conditions, the soil-bentonite slurry wall couldbe built in 1 or 2 weeks and gate section in one week. Slurrywall, gate, 7900 ton of granular iron = ~$400,000

E, VC Funneland-Gate NY facility (1995) treats up to 300 TCE, 500 cDCE, 80 VC (ppb). 12’ x 3.5’ reactive section flanked by 15’ sheet

piling on either side.

Installed in 10 d. VOC reduced to MCLs within 1.5 feet of travelthrough reactive media. Cost $250,000 including $30,000 for 45tons of iron.

CE,

(VI),

In situ Used abiotic reagent sodium dithionite at the Hanford site.

Biotic reagent/nutrient either citrate or glucose.

Compared abiotic & biotic methods for controlling redoxpotential to reduce solids in unconfined aquifer.

Fruchter, J.S., Pacific NW Lab& U. S. DOE. PNL-SA-21731(1993)

y

ents

Tracer Bromide considered most appropriate tracer. Pilot system

(3 long x 3 wide x 5.5 deep (m)) installed 11/95 gov.facility in CO. Models indicate ~ 60 cm/d (2’/d) inreactive zone.

Small tracer not detected. Large targets could disrupt flow; timeconsuming. Water table calculations are not accurate. Heat-pulsevelocity meter give suspect directional vectors. In situ velocityprobes are most promising/easy to use.

Focht, R.M., et al., Internat’lContain. Tech. Conf. & Exhib.St. Petersburg, FL, Feb 9-12,pp. 975 (1997)

o, U,

trate

Batch,

Column

Batch (1 wk) & column (70 d) using Fe foam and steelwool to compare removal of As, Se, Mo, U, sulfate, andnitrate. Precipitation and removal mechanisms alsostudied. 

Batch Fe foam removed 100% Se, 86% nitrate, 100% U, 83%As. Steel wool ~ 80% U, 20% As, 70% nitrate. Neither removesulfate or Mo. Column conductivity decreased slightly (0.08 steelwool; 0.09 cm/s foam). Reduction/ precip. Se; adsorption As, U.

Gallegos, T.J., et al., HSRC/WERC Joint Conf. on theEnvironment, May 20, Paper75 (1997)

Batch TCE (20 ppmv) in 3 neck-1 L round bottom reactors, with

US probe inserted in center neck and tip just above 1.0 gFe or Mg, or 50:50 mix, w/ & w/o US. Controls only US,no metal or US.

Higher pH from dechlorination reactions increases deactivationby precipitating metal compounds on active surface.US can stripaway corrosion keeping metals active longer. Another benefit issonication produces H+ ion to stabilize pH.

Geiger, C., et al., 211th ACSMeeting. New Orleans, LA,March 24-28, 36:17-18 (1996)

ted

nds

Batch 10 g 100 mesh iron filings in 40 mL ZHE hypovials, at 2

rpm. Co for CCl4, HCE, PCE were 1630, 3620, 2250 Fg/L,respectively.

t1/2 CCl4, PCE, HCE = 20, 1100, 13 min., respectively.Chloroform, produced by CCl4 was the only breakdown productaccumulating in significant quantities.

Gillham, R.W. & S.F.O'Hannesin, IAH Conf.,Hamilton, Ontario, May 10-13(1992)
ed...

ant Type test Description/Conditions Results Reference

ted

ons

Batch,

Column,

Reac. Wall

Borden,

Ontario

Batch: 10 g stainless steel, Cu, brass, Al, Fe, & Zn in 40mL hypovials w/ 1,1,1-TCA. Next, Fe w/ halo-aliphatics.Batch/columns used wall material. Constructed Bordenwall by driving sheet piling to form 1.6 m x 5.5 m cell.Reactive material 22% Fe grindings and 78% concretesand with 348 sampling points installed within wall.

Batch: 1st order TCA rates, Steel, brass, Cu low rates; Al better;Zn, Fe rapid (t1/2 100 min.). Fe(0) batch: Highly Cl-organics mostrapid. t1/2 0.22 h HCA to 432 h cDCE. Fe mass to solutionvolume ratio important. Rates decline at pH ~ 9. Batch/column:t1/2 15 h TCE, PCE. In situ wall: Avg. max. conc. downstream ofwall ~10% of influent conc. Performance constant over14-months.

Gillham, R.W., et al., HazMatCentral Conference. Chicago,Illinois, March 9-11, pp. 440-453 (1993)

ted

ons

Batch,

Column

10 g 100 mesh iron powder, silica sand, 40 mg/L CaCO3added to 40 mL hypovials

All 14 chlorohydrocarbons except dichloromethane degraded.Rates in column independent of velocity and consistent withbatch tests. When normalized to 1 m2 Fe surface/mL solution: t1/20.013–20 h. 5 to 15 X > natural abiotic degradation.

Gillham, R.W. & S.F.O'Hannesin, Ground Water,32(6):958 (1994)

CE,

Canister,

NJ

Fe canister, NJ site treating up to 15 PCE, 1 cDCE, 0.5mg/L TCE since 11/94. Initial column used Ni plated Fe,site water. Reactor used commercially plated Ni-Fe 7/96.SA 3.1 m2/g (before plating 1.1 m2/g). 2nd columncommercial Ni-Fe.

t1/2 in initial column 10 X (30 to 3 min, TCE) lower than Fe alone(1 to 1.5 versus 24 h RT). Enhanced reactor t1/2 4X > initialcolumn test, but 4X lower than Fe reactor. Longer t1/2 may resultfrom inadequacies in commercial plating process.

Gillham, R.W., et al.,Internat’l Contain. Tech. Conf.& Exhib. St. Petersburg, FL,Feb 9-12, pp. 85 (1997)

Batch Batch 150 Fm, 370 Fm mesh Fe & Fe powder in 40 mL

ZHEs, TCE in DI water, shaken 150 rpm then analyzedfor pH, dissolved Fe and Cl- removal.

Amount TCE degraded directly proportional to dissolved Fe insolution. 2-fold increase in pseudo first order rate constant whenmetal particle size decreased from 370 Fm by factor of 2.5. Foriron SA/V of solution <1000 m-1 TCE degradation rate constantincreased linearly with SA/V ratio.

Gotpagar, J., et al., Environmental Progress, 16:137(1997)

Batch Varied Cr(VI), H+, and SA of iron as well as ionic strength

and mixing rate.

Rate constant 5.45 x 10-5 cm-2 min-1 over wide range ofconditions. 1.33 mol diss. iron for each mole Cr(VI) reduced.

Gould, J. P., Water Res.,16:871 (1982)

Batch HCl treated Fe particles (<10 Fm). K2PdCl6 w/ Fe powder

(0.05% w/w Pd). 20 ppm PCB (1 mL Aroclor 1260 or1254), methanol/water/acetone (1:3:1) w/ 0.05% w/wPd/Fe in vial, amb. temp., capped, shaken. 1 FL samples.

Rapid (few minutes) dechlorination on Pd/Fe surface. Ratedependant on amount Pd/Fe, % w/w Pd(0) on Fe, & % v/v water.Pd/Fe surface can be used repeatedly if acid-washed after 3 to 4uses.

Grittini, C., et al., EST,29:2898-2900 (1995)

CB Batch,

Column

TCE batch ZHEs: 20 g of 40-mesh Fe & Pd (0.05%)-Fe,100 mL TCE (2 mg/L), surfactant (2%, 4%), cosolvent(2%), 30 rpm. PCBs in 5-mL vials, 2 g of 100-200-meshFe-Pd(0.1%). Columns wet-packed w/ Fe or Fe-Pd,sampled after >10 PV at various levels, times, rates.

Batch Pd-Fe: t1/2 TCE ~27.4 min.; PCB ~100 to ~500 min (assurfactant increased). Columns: TCE and PCBs degrade atenhanced rate t1/2 ~1.5 and 6 min due to increased solid tosolution ratio. Fe-Pd filings applicable for ex-situ treatment ofTCE and PCBs in surfactant solutions generated duringsurfactant flushing.

Gu, B., et al., Internat’lContain. Tech. Conf. & Exhib.St. Petersburg, FL, Feb 9-12,pp. 760-766 (1997)

urans Batch Heat mixtures with Cu to enhance catalyzing reactions to

degrade dioxins and other compounds.

Cu catalyzed degradation of PCDD, PCDF, 7 other chlorinatedaromatics at low temp. similar to first observation using fly ash.

Hagenmaier, H., et al., ES&T,21:1085 (1987)

Batch 1.5-2.0 g of elemental iron per 100 mL aqueous sample

containing 0.02 mmoles TCE (25ppm); maintained pH.

With and without citric acid, pH 5.8: rate 5.37 h-1 & 0.85 h-1; t1/27.74 & 48.9 min., resp. Citric acid chelating ligand for Fe2+.

Haitko, D.A.& S.S. Baghel,209th National ACS Meeting,Anaheim, CA, 35:807 (1995)
ed...

nant Type test Description/Conditions Results Reference

Batch,

Column

Determined that C1 to C5 hydrocarbons are formed by thereduction of aqueous CO2 by Fe(0) and product, have ASFdistribution.

Pretreat iron with H2 increased hydrocarbon conc. 140 h untreated Fe 3.8 ±1.2 Fg/L vs. 7.9±2.4 Fg/L using hydrogenationIndicate absorbed H is a reactant in reduction of aqueous CO2.

Hardy, L.I. & R.W. Gillham.209th National ACS Meeting,Anaheim, CA, April 2-7,35:724 (1995)

bon

ation

Batch Reduction of aqueous CO2 by Fe(0). Reaction mechanism

proposed for electroreduction of aqueous CO2 with Nielectrodes, Fe. Anderson-Schultz-Flory (ASF) distrib.

10 HCs # C5 products ASF distrib. w/ hydrophobics sorbed toFe. Fe supply e- & catalyst promotes formation/ growth of HCchains. H2O also reactant. HC desorption may be rate-limitingstep.

Hardy, L.I. & R.W. Gillham,ES&T, 30:57-65 (1996)

Batch Determine if adding sulfur enhances degradation of CCl4

by iron.

Sulphur (sulfate, organosulphonic acid, sulfides & pyrite) accelerated Fe induced degradation of CCl4 under aerobic conditions. Harms, S., et al. 209th National ACS Meeting, Anaheim, CA, April 2-7,

35:825 (1995)

CE Batch 2 grades Fe, TCE and PCE. Lab grade Fe filings, 420 Fm, 

sulfur >180 ppm. Extra pure Fe, at a particle size of 6 to9 Fm and S content of 22.1 ppb.

Lab grade Fe resulted in ethyne, ethene, ethane in 24 h. Extrapure unreactive after 1 mo. despite high SA, but fast productionof ethyne, ethene, ethane after adding sodium hydrogen sulfide.

Hassan, S.M., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:735 (1995)

enes

Column Columns to assess abiotic/biotic processes in reactive Fe

& Fe-reducing bacteria. Nitrobenzene to analines.

NAC reduction primarily by a reaction w/ surface-bound Fespecies, which serve as mediators for the transfer of eoriginating from microbial oxidation of organic material by Fereducing bacteria-Regeneration of reactive sites, (not e- transferto the NAC) rate-limiting process. Presence of reducible organicpollutants may significantly enhance the activity of Fe-reducingbacteria, in that reduction of such compounds continuouslyregenerates easily available Fe(III) species.

Heijman, C.G., et al., ES&T,29:775 (1995)

Batch CCl4 1.5-5.5 FM; Fe(0) powder: 1 to 10 g per 265 mL

distilled water in anoxic and oxic batch reactors

Products: CHCl3, CH2Cl2. Anoxic Rate: 0.290 h-1, 1 g Fe(0);1.723 h-1, 10 g Fe(0); increases with SA (2.4 mg/g) & time.Slower oxic rates: 0.085 h-1, 1 g Fe(0); DOi 7.4 mg/L. pH rapidincrease after O2 depleted.

Helland, B., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:732 (1995)

-2-

, CT,

zene

Batch 2-chloro-2-propen-1-ol, CT, Chlorobenzene, both bare

and palladized graphite electrodes in aqueous solutionover 24 h. Chlorobenzene rate limiting step indechlorination of compounds such as PCBs.

Pd reduces CT by factor of 5; Both C and Pd/C cathodes effectdechlorination, however Pd/C electrode has a greater selectivityfor dechlorination more products dechlorinated and gives muchhigher yields of fully dechlorinated products. Favor acidic mediaat higher ionic strength, O2 does not affect. 2-chloro-2-propen-1-ol dechlorinates rapidly with Pd/C but by differentmechanism. Chlorobenzene →benzene w/ Pd/C. 

Helvenston, M.C., et al., 213thNational ACS Meeting, SanFrancisco, CA. 37:294-297 (1997)

Column TCE pumped through stainless steel column w/ mixtures

of granular iron & pyrite. Products not considered.

Interaction of GW & geochemical environ. using MINTEQA2.Pyrite in granulated Fe mixtures provided of pH control.

Holser, R.A., et al., 209thNat’l ACS Meeting, Anaheim,CA, April 2-7, 35:778 (1995) 
ed...

ant Type test Description/Conditions Results Reference

Batch Low hexavalent, high hexavalent soils had 105, 460

mg/kg Cr(VI); 1.8, 104 g total Cr; 8.5, 10.4 pH,respectively.

Mn2+ reduced 50 to 100% Cr(VI) in both soils (no pH adjust.).Fe2+ , steel wool reduced soluble & insol. Cr(VI) reductiondependent on pH, reducing agent and soil.

James, B.R., J. of Environ.Quality, 23:227 (1994)

 TCA Column,

reactor,

Belfast,

Ireland

EnviroMetal provided a treatability study w/ Fe filing, sitewater. RT ~ 12 h to reach regulatory limits. Designed 12m tall x 1.2 m dia. Fe reactor with 5 m flow path entry andexit zones to collect and disperse flow. Reactor in cut-offwall to funnel flow.

t1/2 TCE ~ 1.2 h. Small amount cDCE, Cl- increased, VC formedfrom dechlorination (Up to 700 Fg/L) in treatability study.Solvents degraded rapidly. Site Installation: After 7 mo. TCE 5

Fg/L but slightly higher at exit due to backflow from sampling.

DCE formed but ND (not detected) at 3 mo. VC at 4 mo 0.4 Fg/Land ND 7 mo. At 6 mo TCE 2Fg/L and DCE is ND.

Jefferis, S.A., et al., Internat’lContain. Tech. Conf. & Exhib.St. Petersburg, FL, Feb 9-12,pp. 817-826 (1997)

Column 15-cm-dia Plexiglass pipe 90 cm long. Holes every 2.5 cm

first 40 cm, rest 5 cm. Up & downgradient sand zone w/intermed. Fe zone to simulate permeable iron barrier.

CCl4 (up to 1.6 mM) fully dehalogenated by first sample port inFe zone. CCl4 → CHCl3→ CH2Cl2. CCl4 t1/2 occurred within0.25 h at a rate of 2.5 cm/h. CHCl3 t1/2 is slower but increased 2-fold with a 5-fold increase in flow velocity.

Johnson, T.L. & P.G.Tratnyek. 33rd Hanford Symp.on Health & the Environ.—InSitu Remed., pp. 931 (1994)

Column,

Batch

100-mesh untreated Fisher Fe powder (SA 0.057 m2/g)<325 mesh in N2 purged unbuffered water and 20-32 meshFluka Fe turnings SA 0.019 m2 g-1 in carbonate buffer.Mixed in dark, 36 rpm, 23±1EC. Table of new andprevious experimental conditions provided.

New/previous ko sb from batch and column varied widely.Normalization to Fe surface conc. yields specific rate constantkSA (vary by only 1 order of mag.). Dechlorination more rapid insaturated carbon centers and high degrees of halogenation favorrapid reduction. Representative kSA values provided for solvents. 

Johnson, T.L., et al., ES&T,30:2634 (1996)

al

r

Batch,

Column

Hanford Site: Column/batch studies looking at injectionof micrometer-sized Fe(0) colloids into subsurface to formchemical barrier.

Surfactants in low ionic strength solutions increased length oftime dense colloids (7.8 g cm-3 ) remained in suspension by250%. Removal effic. sand column partially controlled byinjection rate.

Kaplan, D., et al., In Si tuRemed., Battelle Press, pp. 821(1994)

al

r

Column PVC column study evaluating Fe(0) colloid injection rate

and concentration on colloid retention by a sand bed.CaCl2 tracer studies to compare transport rates of colloids.

Colloids controlled by rate & influent conc. As colloids accum.,efficiency decreased due to gravitational settling. Colloids wereevenly distributed & high flow required to mobilize.

Kaplan, D., et al., J. ofEnviron. Qual., 25(5):1086(1996)

ub

d

enes 

Batch Suspensions of magnetite, goethite, lepidocrocite,

aluminum oxide, amorphous silica, titanium dioxide inpresence and absence of Fe2+ addition.

Fast nitroaromatic reduction in all Fe hydroxides at 6.5 pH.Mineral oxides (no Fe2+ ) show slow reduction, but increased w/Fe (hydr)oxide coatings. Rates pH-dependent, decreasing w/increasing compound to solids concentration ratio. 

Klausen, J., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:716 (1995) 

ted

ants

Batch Bimetallic process evaluated as a means of increasing

rates of reaction.

TCE dehalogenation increased 2X using Pd/Fe instead of Fe andextends process to less reactive dichloromethane.

Korte, N., et al., 209th Nat’lACS Meeting, Anaheim, CA,April 2-7, 35:752 (1995)

Batch 1 FM CCl4 reacted with 1.2-1.4 m2/L pretreated pyrite at

pH 6.5, 25EC except experiments conducted with sulfideat pH 7.75; aerobic & anaerobic. Pyrite 75-300 Fm.

>90% CCl4 transform in 12-36 d (all conditions). Aerobic >70%CCl4→ CO2 . Anaerobic 50% CCl4 → CHCl3. FeOOH coat onpyrite (aerobic). Pyrite depleted of ferrous Fe in all reactions.

Kriegman-King, M.R . & M.Reinhard, ES&T, 28:692(1994)
ed...

ant Type test Description/Conditions Results Reference

Batch Biotite, vermiculite, muscovite wet-ground to 200-50 

mesh (73-300 Fm). Ampules w/ 13.5 mL buffer, spikedwith solution saturated with CCl4 and flame sealed. 

~ 80-85% CCl4 → CO2 via intermed. CS2. Chloroform 5-15%,5% unidentified nonvolatile and CO. At 25E 1mM HS t1/2 2600,160, 50 d for homogeneous, vermiculite, biotite system,respectively.

Kriegman-King, M.R . & M.Reinhard, ES&T, 26:2198(1992)

Batch CCl4 in vials w/ corrin (B12, cobinamide dicyanide, or

aquocobalamin), reductant (Ti(III), dithiothrietiol, orS2/cystenine), pH 8.2. Products in headspace and mixturedetermined by GC/MS, HPLC, NMR or TLC.

Proposed pathway: trichlromethyl radical forms adduct withreductant. In S2-/cysteine produces CS2 or thiazolidines by way ofthiophosgene. Or radical further reduced to form CHCl3 andCH2Cl2 or CO and formate by dichlorocarbene intermediation. 

Lewis, T.A., et al., ES&T,30:292 (1996)

Batch TCE dechlorination by Zn(0) in aqueous solutions at room

temp. 

Bimetallics Ag, Ni, Pd to enhance Zn. Dechlorination few h toseveral d. Best rates w/ cryo-Zn (ultrafine Zn) & Pd. Ethylene,ethane, monochlorinated hydrocarbons products.

Li, W. & K.J. Klabunde,HSRC/WERC Joint Conf. onEnviron., 5/20 Paper 35 (1997)

Batch ZHEs w/ 3 Fe filings (Fisher, Columbus Chemical,

MBS). TCE 0.5 to 20 ppm. Fisher Fe pretreated withHCl.

Rates varied by factor of 2 for 3 Fe’s. TCE sorbed then reducedby MBS. Fisher Fe 7 to 5 h. TCE/Fe ratios changed rate as well.

Liang, L., et al., 209th Nat’lACS Meeting, Anaheim, CA,April 2-7, 35:728 (1995)

Batch 25g of 40-mesh Fe filings added to ZHEs, containing 125

mL solution (nominally 2 mg/L of TCE) at 30 rpm. Pd/Feprepared according to Muftikian et al. (1995).

When 5 ppm TCE reacts with Fe(0), ~ 140 ppb VC persists 73 d. VC (~ 10 ppb) remaining with Pd/Fe about an order ofmagnitude > w/ Fe(0). Volatile byproducts may be underrepresented in other published data regarding reduction withFe(0). Reduction of TCE w/ Pd/Fe (0.05 % Pd) > order ofmagnitude faster than with Fe(0). With a 5:1 solution-to-solidratio TCE t1/2 with Fe(0) =7.41 h, t1/2 with Pd/Fe = 0.59 h.

Liang, L., et al., GWMRWinter, 122 (1997)

Batch Sodium sulfate, sodium sulfide, ferrous sulfide, pyrite,

organic acid, electrolytic Fe(0) powder, Fe(0) degradationof CCl4 under aerobic conditions.

Products: chloroform, potential for carbon disulphide (toxic).Sulphur significantly increased rates under aerobic conditions.Pyrite can regenerate ferrous ions, produce sulfate & control pH.

Lipczynska-Kochany, E., et al, Chemosphere, 29(7):1477(1994)

ion Batch Tracers, aqueous inorganic profiles, scanning electron

microscopy (SEM), x-ray photoelectron spectroscopy(XPS), and wavelength dispersive spectroscopy (WDS) todetermine precipitates and porosity loss in Fe systems.

Precipitation changes color from black to gray. Loss of alkalinityand calcium, no signif. magnesium loss. Most loss (5-15%) earlyand levels off. SEM shows crystals form on the surface of Fe.Tracers indicate fairly uniform loss of porosity throughout.

MacKenzie, P.D., et al.,Emerg. Technol. in Haz.WasteManag. VII, Sept. 17-20,Atlanta, GA, pp. 59-62 (1995)

CE Column 148 lbs VWR coarse Fe filings in columns. Initially used

buffered DI water (40 mg/L CaCO3). pH 7-8.5. Later usedGW, 400 mg/L CaCO3 and pH 7-8. Flow velocities muchhigher than typical to accelerate effects of aging.

TCE t1/2 36 min at 40, 20, 12 mL/min. tDCE t1/2 `~ 100 min., 1,1-DCE t1/2 200 min, although 1st order fit not as good as for TCE.cDCE particularly poor. Siderite formed at the top of column 1,throughout column 2 and at the bottom of column 3.

Mackenzie, P.D., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:796 (1995)

ion Batch Tracers, aqueous inorganic profiles, SEM, XPS, and WDS

to determine precipitates and porosity loss in Fe systems.

GW forms precipitates (Fe(OH)2, FeCO3, CaCO3) on Fe surfaces,which may affect reactivity. However, this effect, to date, small.Also, H2 produced from anaerobic corrosion of Fe a factorcontrolling the measured porosity losses in iron systems.

Mackenzie, P.D., et al.,National ACS Meeting,37:154-157 (1997)
ed...

ant Type test Description/Conditions Results Reference

Batch, Site

Samples

Three sediment samples from Kesterson, Merced County,CA with elevated levels of Se subjected to in situ Fe(II)amendment.

Both Se(IV) & Se(VI) occluded within FeOOH produced duringFe(II) oxidation & hydrolysis. Fe(II) salt amendment potential insitu remediation for trace Se.

Manning, B.A. & R.G. Burau,ES&T. 29(10):2639 (1995)

Column Rocky Flats seep w/ Cl- organics, metals, radionuclides.

Seep water in glass columns w/ steel wool (Fe(0)) at 5 (1std); 10 (2nd d); 30 mL/min (next 4 d). Rates increased todetermine effects on U removal. DO 5 to 6 mg/L, pH 8,and 13 to 21EC.

Fast removal of U by Fe(0). Chloro-organics determine RT indesign. Studies also test various media on removal of plutoniumand americium at site and excess Fe from barrier effluent.EnviroMetal Tech. Inc. to determine RT for barrier design.

Marozas, D.C., et al., Internat’lContain. Tech. Conf. & Exhib.St. Petersburg, FL, Feb 9-12,pp. 1029-1035 (1997)

Cl4 Batch Evaluated core samples from actual field sites.

Microbiology and geochemistry characterized. 

Granular Fe dechlorinated CCl4 & chloroform. Direct electrolytic process with metal surface may occur in combination withreactions involving hydrogen, ferrous iron, sulfur and microbes.

Matheson, L.J. & P.G.Tratnyek. 1993. 205th ACSMeeting, 33:3

Batch Fe(0) participate by direct reduction, ferrous iron, and

hydrogen produced during corrosion.

Surface reaction predominant in granular Fe. Investigatingpossible microbial activity that might affect dechlorination.

Matheson, L.J. & P.GTratnyek, ES&T 28:2045-2053(1994)

CE,

Batch Temperature, steric, pH dependance of degradation and

reactions of pollutants in response to untreated ironpowder (finer than 100 mesh), under aerobic conditions. 

Electronic (orbital) & conformational preference. Nature ofreductant appeared to determine the stereochemical course ofredox reaction. More investigation needed.

Milburn, R., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:822 (1995)

Column 10 cm dia acrylic pipe w/ ~1,250 mL sand mixed with

test material. U and Mo measured. Used hydrated lime. 

Chemical barriers low cost alternative. Site evaluation todetermine type of chemical barrier (e.g., sorption orprecipitation).

Morrison, S.J. & R.R.Spangler, Env. Progr.,12(3):175 (1993)

Batch Lab experiments to evaluate material for use in chemical

barrier under a repository containing uranium milltailings.

No additive extracted both U & Mo over pH range. Hydratedlime lowered U in pore fluid. Soluble ferrous materials extractedMo.

Morrison, S.J. & R.R.Spangler, ES&T, 26(10):1922(1992)

E, cis

2-di

lene,

Batch Sealed 12 mL glass vial. Pd/Fe [3.6 g-10Fm (Aldrich), or

3.6 g Fe filings (Baker & Adamson), or 10 g- 40 mesh Fe(Fisher)] with 10 mL of chlorinated compounds (20 ppmin H2O), shaken. Sampled with a syringe for GC analysis.

Dechlorination 1,1,2-TCE, 1,1-DCE, cis & trans-1,2-DCE &PCE at 20 ppm to ethane in few min. No intermediate productsdetected at > 1 ppm. Chloromethanes, CCl4, CHCl3, CH2Cl2 alsodechlorinated to methane. CCl4 in a few minutes, CHCl3 in < 1 h,CH2Cl2 in 4-5 h.

Muftikian, R., et al., WaterResearch, 29:2434-2439(1995)

Batch 1x1-cm pure Fe foil, 0.254 mm thick welded to stainless

steel stub etched w/ 8 keV argon ions. Potassiumhexachloropalladate added and allowed to react.

Pd(IV) to Pd(II), protons on hydroxylated Fe oxide form Pd(II)-O-Fe bonds, collapsing to Pd/Fe. TCE forms hydroxylated Feoxide film that deactivates Pd/Fe. Dilute acid removes film.

Muftikian, R., et al., ES&T,30:3593-3596 (1996)

nols

Batch Batch Fe(0)-Pd(0) in unbuffered, DI water, room temp,

dark, usually with [Fe(0)-Pd(0)]= 69.4 g/L and initialaqueous chlorophenols [CP] ~ 0.08 mM, 40 rpm.

Initial rapid loss CP (t1/2# 0.2 h) due to sorption to Fe(0)-Pd(0)surface. Rate constant, kobs, proportional increase with Pd(0) usedand SA of Fe(0)-Pd(0). Higher cosolvent corresponded todecrease in kobs.

Neurath, S. K., et al., 213thASC National Meeting, SanFrancisco, CA, 37:159-161(1997)
ed...

ant Type test Description/Conditions Results Reference

M,

CE

Batch Fe catalyst and aquifer material collected from Canadian

Forces Base, Borden, Ontario. Ci CCl4-4050, TCM-4650,TCE-4080, & PCE-3970 Fg/L at 12EC.

t1/2 2.2, 850, 1520 and 4000 minutes for CCl4, TCM, TCE, andPCE, respectively. No Eh change in controls, reactive vialsshowed highly reducing conditions, but no significant pHchange.

O'Hannesin, S.F. & R.W.Gillham, 45th CanadianGeotechnical Soc. Conf., Oct25-28 (1992)

anic

nds

Wall,

Borden

Reactive wall 22% Fe, 78% concrete sand; 5.5 mdowngradient. Cell driven 9.7 m to bottom silty clay lens.

TCE reduced 95%, PCE 91%. No TCM downstream of wall. Clincrease consistent with quantity degraded. Traces DCE; no VC.

CE Reac.

Wall,

Borden

Field demonstration in 1991 at Borden, Ontario. TCE 270and PCE 43 mg/L. PRB 1.5m wide of 22% granular Feand 78% sand placed in path of plume moving 19 cm/d.

90% TCE and 88% PCE removed from solution and Cl- indicateddechlorination. Major product was cDCE with peak of 2200 Fg/Lfollowed by tDCE and 1,1-DCE, VC ND.

O'Hannesin, S.F., et al.,Emerg. Technol. in Haz. WasteMangmt VII, ACS, Sep 17-20,Atlanta, GA, pp. 55-58 (1995)

Column Plexiglass columns packed with mixture 15% electrolytic

iron and 85% silica sand. TCE pumped at 0.1 mL/min. CiTCE = 1.3, 4.7, 10.2, 61 mg/L. 

TCE Ci = 4.7 mg/L. Products: ethene 40%, ethane 18%, C1 to C410%. 3-DCE isomers & VC. c1,2-DCE primary product ofdegradation, though sum of all chlorinated was only 3 to 3.5%.

Orth, W.S. & R.W. Gillham,209th National ACS Meeting,Anaheim, CA, April 2-7,35:815 (1995)

Column Columns used simulated GW containing 1.3 to 61 mg/L

TCE. Column packed with 15% iron powder mixed withsieved 35-mesh silica sand. SA of iron 0.287m2/g, iron SAto solution volume ratio 0.21m2/mL.

Pseudo-1st-order rates. Products :ethene > ethane >>> otherC1—C4 hydrocarbons. 3.0-3.5% TCE appeared as chlorinatedproducts. Most TCE probably sorbs to iron surface untilcomplete dechlorination achieved.

Orth, W.S. & R.W. Gillham,ES&T, 30:66-71 (1996)

Batch Chlorine solutions stirred in 250 mL reactor at 20EC.

Granular Fe 0.2~.5 dia. added. 10 mL taken at 5, 10, 15,20, 25 min and analyzed for chlorine & chloride contents.Experiments also carried out at pH 4, 5, 6, 7, 8, and 9.

Optimum conditions investigated for pH, particle size, andcontact period. Species of OCl- and HOCl removed by 100%between pH 4 and 7 within 25 min.

Özdemir, M. & M. Tüfekci,Water Research, 31:343-345(1997)

Batch,

Column

Geochemical effects and minerology on reduction CrO42-by Fe(0) using stirred batch reactor (SBRs) and shakenbatch bottles under N2 to evaluate kinetics/mechanisms.

Chromate reduction rapid and complete in zero valent systemsand natural aquifer material. t1/2 = 11 h. When no aquifer materialmuch slower and incomplete during 146 h.

Powell, R.M., M.S. Thesis,University of Oklahoma,Norman, OK (1994)

Batch,

Column

SBRs: under N2 w/ & w/o Elizabeth City aquifer material.C0 of CrO42- = 136 to 156 mg/L. 1st column portion ofaquifer material. 2nd column mixture of Fe filings (7.5 g)& aquifer material (67.6 g). 6 mg/L chromate as K2CrO4introduced at 0.05 mL/min.

Rapid changes in Eh from positive to highly negative uponintroduction of Fe metal. Chromate reduction slow in systemwith no aquifer material but rapid in system containing thenatural solid phase. Eh and pH changes less dramatic in effluentfrom column but influent chromate effectively removed.

Powell, R.M., et al., WaterEnviron. Fed. Conf., March,Miami, pp. 485 (1994)

Batch Some types of iron and aquifer material more reactive

than others. Ada Iron & Metal (AI&M) & MasterBuilders Supply (MBS); Eliz. City, NC & Otis AFB, MAaquifer material used along with commercial Si sand.

Data support conclusion that CrO42- can be reduced to Cr(III) inpresence of elemental iron. AI&M and Eliz City aquifer materialmost reactive. Suitable e- acceptors need to form appropriatecouple. Mechanisms proposed. 

Powell, R.M., et al., ES&T,29(8):1913 (1995)

Batch Aquifer materials from Elizabeth City, NC and Otis AFB,

MA also kaolinite and montmorillonite using simulatedGW. Scrap iron filings from AI&M and cast iron metalchips from MBS.

Coupled corrosion processes responsible for CrO42- reduction &precipitation, w/ AI&M Fe; much greater rates than MBS Fe.Aluminosilicate dissolution proposed to increase reaction rates.

Powell, R.M., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:732 (1995)
ed...

ant Type test Description/Conditions Results Reference

CE Batch Shaken batch bottle experiments were used to evaluate

both the dissolution of 7 aluminosilicate minerals in thepresence of Fe(0) and whether the dissolution affected therates of chromate reduction by Fe(0).

Support previous hypotheses that aluminosilicate dissolutionpromoted Fe corrosion reactions, hence chromate reduction, dueto generation of protons. Proposed mechanisms for chromatereduction and TCE dechlorination depicted in reaction diagrams.

Powell, R. M. & R. W. Puls, ES&T, 31:2244 (1997)

Column Chromate reduction in Fe and quartz grains to determine

fate of reduced Cr in wall material. Fe filings reacted with20 mg/L Cr(VI), as K2CrO7 for more than 150 PV. Quartzgrains flushed w/ CaCO3 solution following Cr(VI)breakthrough in the column.

Reacted Fe developed coatings of goethite with Cr(III)concentrated on outermost edges. In regions of increased Cr,goethite acquired characteristics similar to Fe2O3 and Cr2O3.Complete reduction of Cr(VI) to Cr(III). Cr(III) incorporated intosparingly soluble solid species.

Pratt, A.R., et al., ES&T31:2492-2498 (1997)

e Batch Thiocyanate reacts w/ Fe(III) giving red color (463 nm).

4.5 x 10-5 M Fe2+ (w/ light) accelerates atrazinedecomposition.

Atrazine fully decomposed at pH 1.5 under sunlight within 2 h.Most important factors under light are surface area and nature ofiron used. No degradation in dark.

Pulgarin, C., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:767 (1995)

Char.

Study, Eliz

City NC

48 Cores at various depths. Assessed chemical speciationand distribution of Cr on contaminated soils and itsleaching potential. Batch adsorption/reduction proceduresused.

Adsorption and reduction capacity of soils were overwhelmedpermitting passage of Cr(VI). Capacity differences related to claycontent & pH; less to amorphous iron oxide coating.

Puls, R.W., C.J. Paul, D.A.Clark, J. Vardy.. J. of SoilContam. 3(2):203 1994

CE,

C

Field

Study, Eliz

City NC

25% each by vol. EC aquifer material, sand, Fe-latheturnings (0.1-2 mm) & MBS-Fe-chips (1-10 mm >sulphur, carbon) in field test. 21 augered boreholes, 3 to 8m bgs, were filled with the mixture. 21 monitoring wellsinstalled.

Disappearance of contaminants with appearance of ferrous Fe,decrease in oxidation-reduction potential and DO with slight pHincrease. Sulfide also detected downgradient and within 30 cm ofiron cylinders. Less reducing conditions downgradient.

Puls, R.W., et al, 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:788 (1995)

Field

Study, Eliz

City NC

At Elizabeth City, 2 Fe sources (AI&M, MBS) mixed withnative aquifer material and 10-mesh washed sand. 20-cmdia. cylinders installed in three rows 3 to 8 m bgs, 21total.

Cr to < 0.01 mg/L Significant reductions in TCE. Siderite notdetected, but Fe sulfides were. Full demonstration scheduledJune 1996 of 50-long x 8-deep x 0.6-wide(m) trench of Fe(0).

Puls, R., et al., 4th Great LakesGeotech. & Geoenviron. Conf.,Univ. of Illinois, pp. 23 (1996)

trite Batch Batch in unbuffered, anaerobic, DI water, 22EC, dark, HCl

treated & untreated Fe(0)(>40 mesh) 69.4 g/L. Injectnitrate or nitrite[Co ~ 0.16 mM), 40 rpm. IC for NO3- andNO2- quantification, estimated NH4- using a colormetricHach kit.

1st-order nitrate rate constants, k1 , increased with HCl pretreatedFe(0). First 12 h following treatment, k1 gradually declines in thepresence of Cl-. High chloride in otherwise identical systemscause much smaller decline in k1. Rate constant for nitritereduction, k2, small due to similar acid pretreatment of Fe. k1 &k2 for nitrate & nitrite reduction by untreated Fe(0) directlydependent on concentration of Fe(0), ranging 69.4 to 20.8 g L-1.

Rahman, A. & A. Agrawal,213th National ACS Meeting,37:157-159 (1997)

Batch Electrolytic Fe 5 g/20 mL used. Influenced by preparation

of metal surface by treating with HCl, PCP concentration,pH, temperature, and presence of inorganic compounds.

Pretreated iron improved rates (6 h, 60-70% PCP [2.7 x 10-6 M]degraded). Keep pH near neutral. Some anions (e.g., Cl-) retarddegradation. Results indicated poor remedial choice for PCP.

Ravary, C. & E. LipczynskaKochany, 209th National ACSMeeting, Anaheim, CA, April2-7, 35:738 (1995)
ed...

nant Type test Description/Conditions Results Reference

ion Batch Fe(0) anaerobic corrosion produce Fe2+, OH-, H2(g).

Corrosion rates are measured by monitoring the H2pressure increase in sealed cells containing iron granulesand water. The principal interference is hydrogen entryand entrapment by the iron. The entry rate is described bySievert's law (R= kPH2O5), and the rate constant, k, isevaluated by reducing the cell pressure once during atest.MBS Fe, Blend A, 50/50 mix of 10-18 & 18-32 meshgranules (C, 3.2%; Mn, 0.65%; S, 0.09%; SA, 1.5 m2g-1 ).

For the 10 - 32 mesh iron, k was 0.015 initially but decreased to0.009 mmol kg-1 d-1 kPa-0.5 in 150 d. Corrosion rate in saline GW0.7 ~ 0.05 mmol of Fe kg-1d-1 at 25 °C—identical under saturatedor fully-drained conditions. Rates decreased by 50% in 150 d dueto alteration product buildup. First 40 - 200 h progressivelyincreasing rates of pressure increase. Time before steady-staterates develop depends on the solution composition. Discardedthis data in calculating corrosion rates. Tests on pure NaClsolutions at identical equivalent concentrations (0.02 equiv/L)show the following anion effect on corrosion rate:HCO3>SO42->Cl-. For NaCl solutions, corrosion rates decreasefrom 0.02 to 3.0 mmol kg-1 d-1 kPa-0.5.

Reardon, E.J., ES&T,29:2936-2945 (1995)

CE Site demo,

Moffett,

CA

Pilot field demo, Moffett Field, Mountain View, CA 1/96.50’ long x 10’ wide x 22’ thick funnel- &-gate installedacross TCE and PCE plume 4/96. 

Baseline sampling 6/96 & 9/96 positive. TCE > 1,000 Fg/Lupgradient reduced to ND w/in first 2’ of cell (gate). Democontinue until 3/98; Report to be prepared for DOD.

Naval Facilities Eng. ServiceCenter, Enviro. Restor. Div.,www.Updated April 24, (1997)

Batch,

column

5 to 20 mg/L TCE, 20-kHz US, 0.16 cm2 in 0.5-L bagwith 100-mesh Fe, anaerobic, 160 shakes/min. 

Column: 20% Fe, 80% sand. 4 Fe’s: 50-mesh particles,

Peerless acid-washed chips, unwashed Peerless, MBSwashed chips. 43% 20-mesh; 40% 40-mesh, rest Fe dust.15 mg/L TCE at 4.7 mL/min. US 15.9 mm-dia, auger drillbit, 15 cm long in 50-mesh Fe, 50% power.

US removes corrosion from Fe surface and prolongs reactive lifeSonication for 0.5 h increased rates about 12%. But rates nearlytripled to 184% after 1 h treatment. Prior to US, lower halfcolumn (highest TCE conc.) t1/2 1.5 times upper section. AfterUS t1/2 dropped. Lower t1/2 decrease 70%; upper t1/2 22%.

Reinhart, D.R., et al., Internat’lContain. Tech. Conf. & Exhib.St. Petersburg, FL, Feb 9-12,pp. 806-813 (1997)

DCE Batch 25-mL sealed bottles w/ 1.7 g Fe(0) & 34 mg ground

pyrite in 1.7 mL Ar-sparged DI water. SA Fe(0) 0.7 m2/gwith either 6.7FL tDCE, 4.4 FL cDCE, 10 FL 1,1-DCE,30FL VC in methanol,4 rpm. Volatiles partition into highheadspace slowing further reaction and enhance accum.

2 categories of reductive dehalogenation: Hydrogenolysis(replace halogen by H+) & reductive elimination (2 halide ionsreleased), both net transfer of 2 e-. Haloethylenes can undergoreductive, 0-elimination to alkynes under environmentalconditions. Evidence of this is involved in reaction ofchloroethylenes with Fe(0).

Roberts, A., et al., ES&T,30:2654 (1996)

DD Batch Sealed 40-mL vials. Fe powder (0.3-3 g) with 20 mL

deoxygenated buffer, pH 7, DDT or DDD dissolved inacetone, and, with and without the presence of nonionicsurfactant Triton X-114 at 250 mg/L. Uncapped reactorsflushed with N2, beads to improve mixing. Closed reactorsshaken at 130 rpm at 20 ± 0.5 EC.

Rates of dechlorination of DDT and DDE were independent ofthe amount of iron, w/ or w/o surfactant. Rates w/ surfactantmuch higher than w/o. Initial 1st-order rate of DDTdechlorination was 1.7 ± 0.4 & 3.0 ± 0.8 d-1 or, normalized bythe specific iron SA, 0.016 ± 0.004 and 0.029 ± 0.008 L m-2/h-1),w/ and w/o surfactant, respectively. Mechanistic modelconstructed that qualitatively fit the observed kinetic data,indicating that the rate of dechlorination of the solid-phase(crystalline) reactants was limited by the rate of dissolution intothe aqueous phase.

Sayles, G.D., et al., ES&T,31:3448-3454 (1997)
ed...

ant Type test Description/Conditions Results Reference

Batch System similar to that used by Matheson and Tratnyek

(1994) except with the range of Fe(0) concentrationextended from 5-31 m2/L to 0.2 - 80 m2/L.

Greater range of Fe(0) SA shows hyperbolic relationship.Concentration up to 80 m2 /L with sharp increase in rateindicating heterogeneous catalysis, electrical double layer orabrasion effects during mixing.

Scherer, M.M. & P.G.Tratnyek, 209th National ACSMeeting, Anaheim, CA, April2-7, 35:805 (1995)

Reduction

kinetics

Linear sweep voltammograms (LSV) of Fe(0) at 3000rpm. Potentials set to avoid H+ evolution at more (-). than-700 mV/SHE and O2 evol. at more (+) than 800 mV/SHE.LSVs w/ & w/o CCl4 . Increased (-) current in CCl4attributed to reduction of CCl4 to CHCl3.

More reducing potential on Fe(0) increases rates of H2O & CCl4reduction. However, H2O becomes increasingly larger portion. Atpotentials more negative than -700 mV/NHE, water reductionlarger portion than CCl4 dechlorination, suggesting morereducing potential would not enhance CCl4 dechlorination rate.

Scherer, M.M., et al., 214thNational ACS Meeting, LasVegas, NV, 37:247-248 (1997)

Reduction

kinetics

Fe(0) from 99.5% pure Fe(0) rod 3.0 mm dia., SA 0.071cm2. Mass transport controlled by polished Fe(0) rotatingdisk electrode. Kinetics of CCl4 dechlorination in pH 8.4buffer at potential where oxide film would not form. 

Cathodic current independent of electrode rotation rate, 1st-orderrate constant (kct = 2.3 x 10-5 cm s-1 ) < estimated rate constant formass transfer to surface. Rate reduction of CCl4 by oxide-freeFe(0) dominated by reaction at metal-water interface.

Scherer, M.M., et al., ES&T,31:2385 (1997)

Batch 4.1 g/L Fe powder in oxygen-free, pH 7 water, 50EC. TCE t1/2 = 20 d using Fe. 1,1,1-trichloroethane, 1,1-

dichloroethylene, tetrachloroethylene transformed using bufferedwater & landfill leachate.

Schreier, C.G. & M. Reinhard,Chemosphere, 29(8):1743(1994)

Batch 0.5 g either 0.5% Pd alumina or 1% Pd on granular carbon

(GAC) and 1% Pd PAC at 400 rpm. Total aqueous phase60 mL, N2 purged, capped, 5 mL H2 added for 15 min topre-reduce Pd. PCE = 1.1 Fmoles in 3.5 FL of methanol. 

5 chloroethylenes including PCE and VC reduced in 10 min by0.5 g of 0.5% Pd on alumina and 0.1 atm H2. Ethane 55-85%,ethene # 5%. Pd on GAC yielded 55% ethane from PCE. PCE t1/2= 9 min for 0.055 Fmole Pd (583 Fg of 1% Pd on PAC). 10%nitrite decreased rate by 50%. O2 greatly decreased all rates.

Schreier, C.G. & M. Reinhard,Chemosphere, 31(6):3475(1995)

ed

es

Batch Glovebox 90% N2 / 10% H2, bottles filled with iron and

HEPES-buffered water (pH 7). Placed in 50EC waterbath,with or without shaking.

The more highly halogenated compounds most reactive (tDCE >TCE > PCE = cDCE = 1,1-DCE) with the exception of VC whichis less reactive than PCE. Product: ethene, ethane. PCE:15-30%;VC: 50% reduction. Ethene/ethane ratio larger for VC. TCEintermediate of PCE. cDCE only intermediate of TCE.

Schreier, C.G. & M. Reinhard,209th National ACS Meeting,Anaheim, CA, April 2-7,35:833 (1995)

ted

es 

C)

Batch 125 mL glass bottle, 60 mL N2 purged water, 0.05-0.5 g

catalyst [5 g either 0.5% Pd-alumina or 1% Pd on GACand 1% Pd PAC or PCI-silica], w/ 1-1.1 Fmoles PCE, 400rpm. 5 mL H2 added at stages; some initial H2 = 0.1 atm.Sampled headspace for ethene, ethane, ethylenes.

Pd-PCI-silica PCE rate 0.034 min-1 (t1/2 =20 min), 65% ethane;2% ethene. Some PCE dehalogenated before double bondreduced. Al pellets, GAC or PAC, PCE 99.99% reduced in 10min.; ethane 65-80%; ethene >5%. Sorption significant oncarbon but did not hinder transformation using Pd-PAC.

Schreier, C.G. & M. Reinhard,209th National ACS Meeting,Anaheim, CA, April 2-7,35:749 (1995)
ed...

ant Type test Description/Conditions Results Reference

tes Column ID precipitates that form on iron surfaces w/ differing

water chemistry. Two column tests performed using 100mesh, 99% pure electrolytic iron. A 120 mg/L CaCO3solution passed through one column and a 40 mg/L KBrsolution through other. Rate was at 0.23 mL/min. RT ~13.3 h. N2 gas passed through 2nd column. Sampled after158 and 166 PV.

CaCO3 treated iron formed whitish gray coating on firstcentimeter of column but KBr treated iron did not display anyvisible precipitates. CaCO3 and FeCO3 phases were only presenton the surface of the iron removed from the influent end of thecolumn treated with a CaCO3 solution. Fe surfaces analyzed fromboth influent and effluent end of the KBr treated iron and theeffluent end of the CaCO3 treated iron indicated presence ofmagnetite (Fe3O4) precipitates.

Schuhmacher, T., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:801 (1997)

CE Batch,

Column

Iron powder for removal of TCA and TCE from wastewater.

50% TCA removal from 4 h to 1 h as temperature rose 20 to50EC. Degradation rates highly sensitive to Fe SA, significantdecline at pH values in excess of 8.0.

Senzaki, T. & Y. Kumagai,Kogyo Yosui, 357:2 (1988);Kogyo Yosui, 369:19 (1989);Senzaki, T., Kogyo Yosui,391:21 (1991)

, Cr,

 Mn,

Zn

Electrochemical

Cell

Electrochemical cell of massive sulfide-graphite rockfrom mine site as cathode, scrap iron as the sacrificialanode and acidic leachate from mine site as electrolyte. 

Cell raised pH of ~41 L leachate from 3.0 to 5.6 with decrease inredox potential from >650 to <300 mV. Iron sulfate precipitateformed with a concomitant lowering of Al, Cd, Co, Cu and Ni.

Shelp, G.S., et al., AppliedGeochemistry, 10:705 (1996)

mo-3-

pane

)

Batch Iron powder and H2/Pd/Al2O3.Palladium used as a catalyst

with H2 gas as the reductant. Looked at both sterile(abiotic) buffered and unbuffered conditions. 

Fe(0) dehalogenated DBCP under sterile abiotic conditionsbuffered & unbuffered; also, Pd w/ H2 gas as reductant in GW.pH had little effect, however, a solution with pH = 9 inhibited thereaction.

Siantar, D.P., et al., 209thNational ACS Meeting,Anaheim, CA, April 2-7,35:745 (1995)

mo-3-

pane

)

Batch Compared Fe(0) and H2/Pd-alumina for DBCP →

propane. 4 g of 100-200 mesh Fe powder in 125 mL glassbottle, 110 mL deox. solution, 10 Fg/L DBCP, anaerobic,400 rpm. MilliQtm (DI) water (pH 7.0) or GW (pH 8.2-8.7), some amended w/ anions and/or buffer (pH 7.0).

Fe(0)in H2O DBCP t1/2 = 2.5 min; t1/2 = 41-77 min in GW. O2,NO3- slow reaction. 60 mg/L nitrate removed in 14 min. DBCPtrans. in min. w/ 75 mL GW, 22.5 mg 1% Pd-alumina. Rate inGW 30% slower compared to Milli-Qtm. Slight inhibition inMilli-Qtm by SO4-, NO3-, Cl- or O2. SO32- >> inhibitory effect.

Siantar, D.P., et al., Wat. Res.30:2315 (1996)

e Batch, Soil Fe(0) (10% w/w), 0.02 mg/L atrazine in batch. 20 mg

14C-atrazine in Fe(0) (20% w/w) in batch. Fe(0) (2% w/w)in soil to determine mineralization and availability of thepesticide atrazine.

The batch test removed 93% atrazine in 48 h. 5% of that wasadsorbed ?readily available”; 33% "restricted"; 2% residues. 88% 14C removed in 48 h of that 6% was available; 72% pool;the rest was bound. Fe in soil quadrupled mineralization in120-d. 2% Fe(0) & 100 mg NO3- kg-1 increased by a factor of 10;unextractable residue was greater than two times greater than thecontrol (no Fe(0)). 

Sinah, J., et al., HSRC/WERCJoint Conference on theEnvironment, May 20, Paper36 (1997)

, VC Batch,

Column

Anaerobic & mildly aerobic conditions; > 25 commercialiron metals in several forms; 0.1-1325 m2/L iron metal;several groundwaters used.

No significant products from TCE batch or column. Strong Fe(0)

B-bonds may prevent DCE & VC products from desorbing.

Direct reduction of adsorbed chloroethene at metal/waterinterface. Reduction by iron oxide and oxyhydroxide not seen.

Sivavec, T.M. & D.P. Horney,209th National ACS Meeting,Anaheim, CA, April 2-7,35:695 (1995)

Batch,

Column

Effect of FeS varying the FeS mass in batch. Columnswith Fe filings, SA concentration avg. 6000 m2/L. 

TCE t1/2 ~40 min. Rates constant over several hundred PV eventhough surface of iron filings coated with FeCO3 precipitate.

Sivavec, T.M. et al. Emerg.Technol. in Haz. Waste Mngt.VII, Atlanta, pp. 42-45 (1995)
ed...

ant Type test Description/Conditions Results Reference

Column Bimetallics accelerate degradation relative to untreated

Fe. Column of Ni-treated granular Fe with TCEcontaminated site GW (TCE 2.1-3.3 mg/L). 

1st-order TCE rates and products in > 250 PV in Fe/Ni. In 76 PVrates accelerated above untreated Fe. Catalytic dehydrohalogenation to hydrogenation caused enhancement. But, decreaseduntil rate similar to untreated iron. Gray precipitate after 100 PVGW (250 mg/L carbonate). Fe catalysts prone to deactivation.Similar losses not shown in granular Fe(0)systems.

Sivavec, T.M., et al. Internat’lContain. Tech. Conf. & Exhib.St. Petersburg, FL, Feb 9-12(1997) 

Sivavec, T.M., et al., 213th

National ACS Meeting,37:83-85 (1997)

Packed Bed Packed bed of Fe sorbent supported on fine mesh stainless

steel screen & Teflon® flakes at 0.12-0.70 cm/s.

Cd(II) 5 mg/L at pH 7, flow 1.6 mL/min. ~ 8,000 bed volume of

synthetic waste treated before breakthrough of Cd(II).

Smith, E. H., Emerg. Technol.

in Haz. Waste Mngmt. VII,Atlanta, GA, pp. 1205 (1995)

Batch

Column

Batch test of test material with synthetic U mill tailing

pore fluid. The column consisted of a solids chamber anda water sampling chamber. 5 bottom chambers filled w/FeSO4 and sand, 5 top chambers w/ lime and sand.

Redox front coincided w/ precipitation of ferrous iron by contact

w/ Ca(OH)2. Mo & U successfully removed for 6 & 9 d, respec.U reduced to UO2 & precipitated as CaUO4 from elevated pH.Mo reduced to Mo3O8 or MoS2, or precipitated as FeMoO4.

Spangler, R.R. & S.J. Morrison,

Pasco, WA, U.S. DOE Report(1991)

ions Model A variety of configurations simulated using FLOWNET

ver. 2.0, a 2-D steady-state flow model based on dualformation of flow.

2-D model shows width of capture zone proportional to discharge

through gate. Most efficient configuration is sides 180E apart,oriented perpendicular to the regional hydraulic gradient.

Starr, R.C.& J.A. Cherry,

Ground Water, 32(2):465(1994)

ted

s

Batch First environmental application for removal of chlorinated

organic compounds from aqueous solution.

Catalyzed metallic iron powder was shown to degrade a wide

range of halogenated organic contaminants.

Sweeny, K.& J. Fischer, Patent

3,640,821 (1972); Patent3,737,384 (1973); Patent4,382,865 (1983)

Field Site,

CA

VOC degradation rates in GW through 7’ Fe canister. 42.9

h RT determined 2.2’ width wall. Steel plates divided~220 tons granular Fe in center from outer pea gravel. 4monitoring wells downgradient; 2 piezometersupgradient.

First commercial in situ iron wall treating VOCs at former

semiconductor facility. Design & construction from 11/94 to1/95. Operation and regulatory issues summarized. Monitoringshows water quality objectives being met.

Szerdy, F.S., et al, ASCE

National Convention, Nov 12-14, Washington, D.C. pp. 245-256 (1996)

es Batch Open batch system of 1000 mL. H2SO4 or NaOH used to

adjust pH. Optimal pH 2-3, well mixed, ratio H2O2 to iron0.001M:1 g/L

H2O2/iron powder is better than the Fenton's reagent due to

continuous dissolution of iron powder and dye adsorption topowder even though Fenton's reaction major decolorizing agent.

Tang, W.Z. & R.Z. Chen,

Chemosphere, 32(5):947-958(1996)

Batch,

Column

250 mL sealed bottles w/ organics, nutrients, pH 7.0,

Shiprock bacteria. Columns w/ straw, alfalfa, sawdust,sand (25% foc). GW 15 mL/ h (1 d RT), 1 PV bacteria. 

Sulfate, nitrate, U(VI) monitored 90 d. Precipitated U(IV)

crystalline UO2(s). Batch and column results support use ofcellulosic substrates as candidate barrier materials.

Thombre, M.S., et al., Internat’l

Contain. Tech. Conf. & Exhib.St. Petersburg, FL, Feb 9-12,pp. 744 (1997)

Batch Steel nuts put in barrel immersed in 2-L soln containing 90

ppm Cr(VI); at 16 rpm.

After 7.2 h, final Cr concentration < 0.5 Fg/mL. Thornton, R.F., Patent

5,380,441 (1995)

mo

e Batch Determine e- transfer during reductive dehalogenation. Evidence from stereospecificity of reductive elimination

of vicinal dihalide stereoisomers synthesized in lab.

meso-2,3-dibromopentene → >95% trans-2-pentene; D,L-2,3-

dibromopentane >95% cis-2-pentene. Reduction at metal surfacewhere 2 e- transferred w/ no free radical intermediate.

Totten, L.A.& A.L. Roberts,

209th National ACS Meeting,Anaheim, CA, April 2-7,35:706 (1995)
ed...

ant Type test Description/Conditions Results Reference

Column 1st column: CCl4 in air-sat. DI water. All O2 consumed by

Fe(0) giving anoxic downgradient region. 2nd similar butFe zone longer and diluted by mixing with sand.

DO reacts with Fe(0) slowing dechlorination. But oxygen fromair accelerated reactions, possibly due to pH effect fromcarbonate, changes in pathway, catalytic role of Fe2+, or O2creating active corrosion sites.

Tratnyek, P.G., et al.,Emerging Technol. in Haz.Waste Manag. VII. Atlanta,pp. 589 (1995)

ted

s

Reactivetransport

model, kSA

Average estimates of rate constants. Assume reductivedechlorination. Deviations with longer exposure due toprecipitates. 1st-order predictions vulnerable to changes inmechanism or rates for less reactive constituents.

Pseudo 1st-order rates normalized to Fe-SA(kSA). Solvents compared over range of conditions. kSA varies by concentration, Fetype, etc. Representative kSA’s and reactive transport modelcalculate minimum barrier width for flow velocities andhalocarbon.

Tratnyek, P. G. et al., GWMR,Fall, pp. 108 (1997)

CE,

VC

Column NJ site GW. Major VOCs- PCE # 50 mg/L, TCE # 3

mg/L. TDS 425-450 mg/L. 100% iron in column.

t1/2 PCE, TCE, cDCE, VC = 0.5, 0.5, 1.5, 1.2 h, respectively. 2ndtest similar for PCE, TCE but cDCE = 3.7 & VC 0.9 h.Corrosion increased pH & promoted precipitation of CaCO3,FeCO3 , FeOH2. 

Vogan, J.L., et al., (1994)

ed

s

Column Column using site water and Single Layer Analytic

Element Model to evaluate treatment zones, flowvelocities, and residence times.

GW flow model & degradation rates to design and estimate costfor full-scale funnel-and-gate system at shallow sand aquifer (30-40 ft) at Army Ammunition Plant, MN.

Vogan, J.L. et al., 87th Ann.Mtg, Air & Waste Manag.Cincinatti, OH, (1994)

CE,

VC

Pilot-scale

Test

NJ site. Above-ground reactor influent PCE of 30 mg/L.RT 1.1 d. Flow rate 0.5 gpm for 3 mo.

Assumptions: Time for PCE degradation sufficient for any TCEto degrade; 10% cDCE, 1% VC from PCE & TCE degradation. 

Vogan, J., et al. 209th NationalACS Meeting, Anaheim, CA,April 2-7, 35:800 (1995)

Batch Kinetics dependent on pH, SA of metal, CCl4 conc., buffer

and solvent composition (volume fraction 2-propanol).

Reduced CCl4 to chloroform in few h. Rate was 1st-order withrespect to CCl4 at concentrations < 7.5 mM.

Warren, K., et al., J. of Haz.Mat., 41:217 (1995)

Surface

chemistry

XPS identified surface elements, valence state. Alfa AesarFe (10 x 1mm) cleaned (H2O2, hydrofluric acid, H2O).Vapor purified PCE, adsorbed to an Fe (100) singlecrystal (10 mm in diameter and 1 mm in thickness)purchased from Alfa Aesar by exposing 1 x 10-7 Torr 100s or 5 x 10-7 Torr 200 s at room temp.

PCE adsorbed to metal surface, activated by chemisorption. Clfrom e- transfer from Fe to adsorbed species. Adsorbed water candissociate and provide H+ for C surface species from PCEdissociation. Hydrocarbon can be produced from this reaction.

Wang, C.-B.& W.-X. Zhang,213th ACS Meeting, SanFrancisco, CA, 37:163-164 (1997)

ste Batch Sealed, glass flasks simulated mine drainage, organic

mixtures, measured permeability. Consortium of bacteriafrom creek facilitated reducing conditions & degradation.Limestone ensured optimum pH; anaerobic conditions.

Reactivity & permeability (> 10-3 cm/s) suitable. Higher sulphatereduction rates and longer effectiveness from organic mixture.Geochemical reactive and transport models will be used to assesseffectiveness in treatment of mine drainage using reactive walls.

Waybrant, K. R., et al,Sudbury '95, Mining and theEnviron. CANMET, Ottawa,Ontario, 3:945-953 (1995)

Batch Anaerobic, 200 rpm, 20EC, 25 mL, methanogenic culture,

iron powder, iron filings, steel wool. CHCl3 tested in ironcell, iron-supernatant and resting cell.

k = 0.11 (Fe-cell), 0.003 (Fe-supernatant) & 0.007 hr-1 (restingcell). Biodehalogenation to abiotic reactions 37:1. Biocorrosionof Fe & biodehalogenation of CHCl3 via cometabolism using H+from H2O.

Weathers, L.J., et al., 209thNat’l ACS Meeting, April 2-7,Anaheim, CA, 35:829 (1995)

ntane Batch Synthesized stereoisomer to demonstrate e- transfer. Experiments indicate reduction takes place at metal surface. Weber, E.J., 209th Nat’l ACS

Anaheim, CA, 35:702 (1995)
ed...

ant Type test Description/Conditions Results Reference

azo

(4-

Batch 4-AAB has reactive amino group for attaching molecule

to nonreactive surface. Reducing azo linkage suggestsaqueous reductant, if not, then surface- mediated process.

Circumvented surface mediated contact to surface of Fe(0) byadding appropriate water soluble e- mediators. Species that canfunction as e- mediators were found present in the soil. 

Weber, E.J., ES&T, 30:716(1996)

TCE,

VC

Treatment

Wall

Granular Fe(0) & gravel at former semiconductor facility,San Francisco. GW flow 1’/d. 2 d RT required for VC.

Wall 40’ long, 20 to 7’ bgs, Fe 4’ wide allow 2 d RT. Slurrywalls east and west side for hydraulic control. 4 mon. wells inwall.

Yamane, C., et al., 209th Nat’lACS Meeting, Anaheim, CA,April 2-7, 35:792 (1995)

e Batch 150 mL flasks in air-dry & sat. (parafilm seal) 3 d.

Samples extracted in 50 mL DI. Purged with N2. 

"Artificial" soil contam. with nitrate (10 g clay + 10 g sand

+ 10 g Fe powder + 10 mL of 50 mg-NO3--N/L). 

94.4% Nitrate removal using 0.01M HEPES & 6% (w/v) Fe(0).Treating 60 mg NO3--N/L w/ 6% Fe(0) at pH 1.0 transformed allin 24 h; nitrate transformation is inversely related to pH. Nitratereduced to ammonium. Optimal removal at Ci NO3--N = 50-80mg/L. In soil up to 97% transformation in air-dry samples, 99%in wet samples; 2% in controls. 

Zawaideh, L.L., et al.,HSRC/WERC Joint Conf. onthe Environment, May 20(1997)

CE Batch 20 mg/L TCE, 1.0 g nano-Fe or Pd/Fe in 50-mL vial

sealed, 30 rpm. PCBs, 50 FL 200 Fg/mL Aroclor 1254 inmethanol w/ 1 mL ethanol/water (1/9), 0.1 g wet Fe orPd/Fe in 2-mL vial, 30 rpm, 17 h. Commercial Fe tested.

Synthesized sub-colloidal metals. Nanoscale Fe more reactivethan commercial Fe powder, due to high SA, active Fe surface,less surface coverage by iron oxide layer. Nanoscale Pd/Fe moreactive than pure Fe but nano-Fe inactivated by Fe oxideformation.

Zhang, W.-X. & C.-B. Wang,213th ACS Meeting, SanFrancisco, CA, 37:78-79(1997)

Batch Pyrite fines collected near mine area. Pyrite crushed to -45

Fm. Used mixed batch reactors.

Pyrite found to act as efficient Cr(VI) reducing agent. The Cr(III)hydroxide precipitated onto pyrite particles.

Zouboulis, A., et al., Wat. Res.29(7):1755 (1995)
74

Appendix B. Explanation of Relevant Physical/Chemical Phenomena

Corrosion

Corrosion processes are the required chemical underpinning of contaminant remediation by metallic iron. Until

recently, the science of corrosion was concerned almost exclusively with studying the process in order to minimize,

and if possible eliminate, corrosion reactions. Corrosion results in rusting of the iron and steel in cars, pipes,

bridges, buildings, and other structures. It is perhaps the most expensive aspect of infrastructure deterioration in

modern society. Although corrosion is familiar to everyone as rusting, few are aware of the extremely complex

reactions occurring within the corroding metals, at their surfaces and, in the case of electrochemical corrosion, in

the surrounding electrolyte solutions. To understand how these reactions can be used for remediating contaminants

requires at least a fundamental awareness of the manner in which corrosion reactions proceed.

Zero-valence-state metals, such as metallic iron (Fe(0), can serve as electron donors for the reduction of oxidized

species (Sculley, 1975). These metals are unstable in the natural environment and have to be created using hightemperature metal refining processes (Evans, 1960; Sculley, 1975; Snoeyink and Jenkins, 1980). Zero-valence-state

metals tend to revert to a form that is more thermodynamically stable; for example, iron metal oxidizes to Fe2

O3 in

the earth’s oxygen-rich atmosphere. At low temperatures the rate of simple atmospheric oxidation of iron and steel

is negligible, however, due to the formation of oxide films that inhibit further surface exposure.

When a metal is immersed in an aqueous salt solution, as would be the case for a reactive barrier of iron chips or

filings in an aquifer, an electrochemical corrosion mechanism will occur. Electrons are given up by the metal in one

area (the anodic region) forming soluble cations of the metal, and taken up by oxidized species that become

reduced, at another part of the metal surface (the cathodic region). The instability of the iron itself can provide the

necessary energy for oxidation-reduction reactions without external energy input, provided suitable coupled

electron-accepting reactions can occur with reducible species at the cathode.

Typically dissolved oxygen is the preferred oxidant, or electron acceptor, during aerobic corrosion processes.

These systems can, however, become anoxic or anaerobic if oxygen is depleted by the reactions. When present,

inorganic contaminants such as chromate (CrO4

2–) or highly halogenated organic compounds such as PCE and TCE

can serve as the oxidants, accept electrons, and become reduced. As long as electron acceptors are present,

corrosion processes and electron transfer within the metal can continue.

An electrochemical corrosion cell (ECC) can form in a number of ways, including:

1) the simple contact of two different metals. One will become the anode, the other the cathode. The

position of the metals in the galvanic series determines the direction of electron flow; i.e., which

becomes the anode and which the cathode.

2) when anodic and cathodic regions develop on the same metal surface. This can result from compositional

variations within the metal (i.e., other metal contaminants or inclusions with differing galvanic

potentials), surface defects, differences in grain structure orientation, stress/strain differences, and

chemical variations in the surrounding electrolyte solution (Evans, 1960; Sculley, 1975; Snoeyink and

Jenkins, 1980; Adamson, 1990).

These electron transfer processes within a metal or between contacting metals are said to occur within the external

circuit of the electrochemical cell. An internal circuit is also required to complete the cell. This requirement is

fulfilled by the contacting electrolyte solution. This electrolyte can consist of water containing salts (such as ground

water) and reducible, (i.e., electron-accepting), solute(s). In some cases the water itself can accept electrons via

reductive dissociation. This is easily observed when the zero-valent metal is very low in the galvanic series, such as

Mg. When Mg metal is added to even deionized water (i.e., no solutes present), bubbles of H2

 gas rapidly appear on

the Mg surfaces. Mg is so low in the galvanic series that the water itself serves as the electron acceptor and is

rapidly dissociated. As you proceed to metals higher in the galvanic series, the dissociation of water becomes less

and less pronounced.

The ECC is basically a low-power version of the same type of circuit that is established in batteries. Figure 18

shows a simple ECC in a beaker of electrolyte solution with the external circuit resulting from the contact of two

dissimilar metals, tin and iron. The electrons travel through the external circuit from the anode (Fe0

) to the cathode

(Sn0

), where they can reduce oxidized species.

B -
75

Some of the potential corrosion and contaminant reduction reactions in zero-valent iron systems are:

Fe0 → Fe2+ + 2e- Anode (25)

2H+

 + 2e- → H2 (gas) Cathode (26)

Fe0 + 2H+ → Fe2+ + H2 (gas) Net reaction (27)

and, when oxygen is present (aerobic system)

4Fe0 → 4Fe3+ + 12e- Anode (28)

12H+

 + 3O2 + 12e- → 6H2O Cathode (29)

4Fe0

 + 3O2 + 12H+ → 4Fe3+ + 6H2O Net reaction (30)

where the increase in pH due to proton consumption (Equation 30) results in the precipitation of the Fe3+ as

Fe(OH)3

. Should chromate be present as an oxidized species

Fe0 → Fe3+ + 3e- Anode (31)

CrO4

2- + 4H2

O + 3e- ↔ Cr(OH)3 + 5OH- Cathode (32)

Fe0 + CrO4

2- + 4H2

O → Fe(OH)3 + Cr(OH)3 + 2 OH- Net reaction (33)

Gillham and O’Hannesin (1994) have proposed that the reductive dechlorination of chlorinated organic compounds

by iron metal corrosion may proceed as

2Fe0 → 2Fe2+ + 4e- (34)

3H2

O ↔ 3H+ + 3OH- (35)

2H+

 + 2e- ↔ H2 (gas) (36)

X-Cl + H+

 + 2e- ↔ X-H + Cl- (37)

2Fe0

 + 3H2O + X-Cl ↔ 2Fe2+ + 3OH- + H2 (gas) + X-H + Cl- (38)

Figure 18. Example of an electrochemical corrosion cell.

B -

C

a

t

h

o

d

e

A

n

o

d

e

External

Circuit 

(contact of

dissim ilar

m etals) eS n

F e

2 H + + 2e - H 2

Internal Circuit 

(electrolyte)

Corrosion Pit

eF e 2+ Cathode Anode
76

Which of these reactions, or others, is dominant will depend on the conditions and contaminants present; for

example, the presence or absence and the concentrations of other reactive species (including the partial pressures of

gases such as O2

 and CO2) and mineral surfaces, pH, etc. It should be noted that these reactions tend to increase the

pH of the corrosion system. It is also important to note that although chromate is reduced and chlorinated

hydrocarbons are reductively dechlorinated in the presence of Fe0

, current research indicates that the mechanisms

of the reactions are different.

Sorption, Adsorption, Dissolution, and Precipitation Processes in Zero-Valent Iron Systems

As a contaminant plume moves through the subsurface, chemical processes occur that can affect both contaminant

concentration and the overall hydrogeochemistry of the system. Three of the most important physical processes are

sorption, dissolution, and precipitation. These processes also occur in iron PRBs, where the radical geochemical

changes can exert profound effects on the ground water and contaminants.

Sorption and adsorption are loosely described as processes wherein chemicals partition from a solution phase into

or onto, respectively, the surfaces of solid phase materials. Both sorption and adsorption at particle surfaces tend to

retard contaminant movement in ground water. Retardation of the contaminant at the iron surface is a positive

result, allowing increased time for reactions.

In the subsurface context, sorption usually implies movement, or dissolution, of a chemical into a surface coating

on an aquifer material mineral grain. For example, hydrophobic organic compounds such as PCE tend to be sorbed

into organic carbon coatings on mineral grains. In fact, evidence seems to indicate that graphitic inclusions/coatings

on some granular iron surfaces might enhance reductive dechlorination relative to iron without such organic

materials being present at its surfaces. It is possible that sorption of the chlorinated compounds to the graphitic

constituents increases the contact time between the contaminant and the iron surface. This would allow increased

reaction time and better proximity for the requisite electron transfers to occur.

Adsorption implies attachment of a chemical to reactive sites on mineral surfaces. These sites usually result from

an excess or either positive or negative charge on the surfaces. These surface charges can be constant (fixed) due to

ion substitutions in the mineral matrix (isomorphous substitution), variable with pH, or a mixture of both. In

addition, the adsorption can result from either inner-sphere or outer-sphere complexation. Inner-sphere complexation is due to actual covalent and ionic chemical bond formation. In outer-sphere complexation adsorption results

from ion-pair bonding due to electrostatic forces and hydration water separates the solvated ion from the surface.

 Many metal oxides and some clay minerals have net surface charges that vary with pH due to the proportion of

protonated versus deprotonated surface sites. Among these variably-charged materials are the iron oxyhydroxides

(rusts) that result when zero-valent iron corrodes. These materials are very significant to adsorption of both

inorganic and organic charged solution species (ionic species). The charges of both the surface and the solution ion

control whether adsorption will occur or whether the surface and the ion will repulse one another. The pHzpc , or pH

of zero-point of charge, is the pH at which negatively and positively charged surface sites exist in approximately

equal numbers on the mineral. Above pHzpc the surface will have a net negative charge, enhancing cation

adsorption; below pHzpc the surface will have a net positive charge, enhancing anion adsorption. Research has

shown that the reduction of the negatively-charged Cr(VI) chromate ion (CrO4

2-) to Cr (III) by zero-valent iron

occurs more rapidly when the system pH is below the pHzpc ≈ 8.5 of the iron oxyhydroxide rusts; i.e., the rusts have

a net positive charge (Powell et al., 1995). This increase in reduction is probably partially due to the contaminant

being maintained at the iron surface by adsorption.

Although flowing subsurface systems are not in true chemical equilibrium, they can establish a hydrogeochemical

pseudo-steady state condition. This condition can be altered when a contaminant front/plume passes through (Puls

and Powell, 1992). Contaminant fronts typically have pH, Eh, ionic strength, chemical species and complexants,

and other features that differ from the intruded pseudo-steady state system. These differences can, among other

things

• dissolve cementation between the mineral grains,

• change the adsorptive nature of the solid phase (potentially causing charge reversal and/or allowing desorption

of previously immobilized species and/or adsorption of previously mobile species),

• alter mineral-bound and aqueous elemental oxidation states, or

• precipitate new phases onto the solid surfaces.

B -
77

The implementation of a reactive iron barrier wall in the subsurface can be viewed as having effects on the aquifer

in the vicinity of the wall that are analogous to a plume moving through the zone. Intra-wall and downgradient

effects will include radically lowered Eh and loss of dissolved oxygen, higher pH, and increases in Fe2+ that will

precipitate downgradient, possibly as iron oxyhydroxide colloids. Loss of cementation and precipitate formation

can generate mobile colloidal particles, that can themselves transport adsorbed contaminants, or contain toxic

materials that were naturally occurring in the geologic matrix. Additional research is needed on these disruptive

effects to determine whether certain types of intercepted contaminants can be transported colloidally or, alternatively, whether naturally-occurring but immobilized metals in or on the aquifer minerals might be liberated and

mobilized by the dramatic geochemical changes.

The dissolution and precipitation processes associated with the iron metal are, however, essential for contaminant

remediation. The reduction of Cr(VI), for example, occurs primarily due to the anodic dissolution of Fe2+ during the

corrosion process (Powell et al., 1995). The subsequent immobilization of the resulting Cr(III) is due to

precipitation of the chromium as hydroxides or mixed iron/chromium oxyhydroxide solid solutions. Although the

chlorinated hydrocarbons are not precipitated during remediation, dissolution of the iron still occurs as Fe(0) yields

electrons to the hydrocarbon and the resultant Fe2+ is exposed to the solution. Understanding the dissolution/

precipitation geochemistry of these zero-valent iron systems and the contacting aquifer materials is an area of

ongoing research.

B -
78

Appendix C. Scoping Calculations

Cost-effective use of permeable reactive barriers for ground-water treatment requires proper estimation of the

amount of reactive material required and choosing the best means of emplacing it in the ground. The weight of

reactive material per unit cross-section of the plume may be estimated from laboratory reaction kinetics data and

basic knowledge of the plume and the remediation goals. The value of this parameter has implications regarding the

choice of permeable barrier design and emplacement method. The use of tremie tubes, trenching machines, highpressure jetting, and deep soil mixing may be appropriate for different situations, depending on the amount of

reactive material required, the dimensions of the plume, and other factors. The specific application considered here

is granular iron to treat ground water contaminated with chlorinated solvents, but the principles are applicable to

other types of media and contaminants.

Reaction rate parameters from laboratory studies of iron-mediated degradation of a variety of chlorinated solvents

have appeared in the literature in the past several years (Johnson et al., 1996; Shoemaker and al., 1996). The work

of Johnson et al. (1996) has been especially helpful in establishing the high degree of consistency between kinetics

data obtained from batch and column studies. By expressing rate data in a way that accounts for the iron surface

area concentration, it was demonstrated that results reported in the literature varied by less than had previously been

thought. This makes it possible to obtain a fairly reliable estimate of the iron requirement for a potential application

even before site specific laboratory feasibility tests are conducted. The bulk of the data reviewed by Johnson et al.

suggest a surface-area-specific rate parameter (kSA) of about 0.2 cm3

 h-1 m-2 for TCE and of about 0.04 cm3 h-1 m-2 for

cDCE. For the examples considered here, 1.0 m2

/g will be used for the specific surface area, a value typical of the

granular irons which currently appear to be the most practical for permeable barrier applications. Further, the rate

of reaction will be decreased by 50% to adjust for subsurface temperatures being lower than room temperature

(Sivavec and Horney, 1995). Therefore, the effective rate parameter to be used is 0.1 cm3

 g-1 h-1 for TCE and

0.02 cm3

 g-1 h-1 for DCE.

Two example cases are considered below. The first and simplest involves degradation of a chlorinated compound

(e.g., TCE) where the levels of intermediate products (e.g., DCE) are low enough that they do not influence the iron

requirement. The second case involves significant generation of an intermediate product that degrades more slowly

than the parent and thereby determines how much iron is required.

Case 1: Parent Products Only

The rate of reaction may be expressed as

dP

dt = − k P 1ρ m (39)

where P is the concentration of dissolved chlorocarbon, t is the contact time between the dissolved chlorocarbon

and iron particles, k1

 is the first-order rate parameter, and ρm is the mass of zero-valent iron particles per solution

volume. This equation may be integrated to give

ln P

P

k t m

0

1

F

H IK = − ρ (40)

where P0

 is the initial concentration of dissolved chlorocarbon. In a batch laboratory experiment, k1 may be derived

from the slope of a semi-log plot of P0

/P vs. time.

For the case of steady-state flow in a packed bed reactor, an expression analogous to Equation 40 may be derived

by expressing the residence time (t) as the product of the bed void fraction (ε) and the reactor volume (V), divided

by the liquid flowrate through the bed, yielding

ln P

P

k V

Q

F 0 1 m

H IK = ρ ε (41)

The term ρmεV is the mass of zero-valent iron, W, that the fluid encounters as it flows through the bed. With this

substitution, and by representing the flowrate as the product of the cross-sectional plume area (A), the soil porosity

(n), and the average flow velocity (u), the amount of iron required per unit cross-section of plume to effect a desired

decrease in chlorocarbon concentration may be expressed as

C -
79

W

A

un

k

P

P = F

H IK 1

0 ln (42)

This is a useful expression because it allows estimates to be made without assuming a particular design (such as

funnel-and-gate) or calculating parameters such as residence time, but rather expresses a key aspect of the design

(W/A) in the most fundamental terms. However, it does not reflect uncertainties and fluctuations in parameter

values that must be considered in any design. These can be accounted for in terms of a factor of safety (F) which

increases the amount of reactive material employed:

W

A

F un

k

P

P = F

H IK 1

0 ln (43)

A Monte-Carlo simulation has been developed to estimate appropriate factors of safety for permeable reactive

barrier systems (Eykholt, 1997). With influent concentrations varying 10%, the reaction rate parameter varying

30%, and the ground-water velocity varying 100%, achieving a 1000-fold decrease in contaminant concentration

with 95% confidence was found to require a safety factor of 3.5.

As shown in Table 9, calculations based on a safety factor of 3.5 and a range of practical values for reaction rate

parameters and ground-water velocities suggest that W/A should be expected to vary from as little as about 20 lb/ft2

to perhaps 1,000 lb/ft2.

Case 2: Significant Intermediate Generation

If significant amounts of intermediate products are generated during the degradation of a parent chlorinated

compound, the slower rate of degradation of the intermediate product may be the factor that determines how much

iron is required (Focht et al., 1996). The kinetics of intermediate product generation and degradation may be

expressed as

dP

dt = − k P 1ρ m (44)

dD

dt = − αρ ρ k Pk D 1 2 m m (45)

Table 9. Required Weight-per-Area (W/A) of Granular Iron. Calculated with n=0.33, P0/P=1000, and F=3.5. Information

and Equations from Appendix C.

k1

(cm3 g-1 h-1)

u

(ft/day)

W/A

(lb/ft2)

0.1 0.1 21

0.25 52

0.5 100

1 210

0.02 0.1 100

0.25 260

0.5 520

1 1000

C -
80

where D is the concentration of the intermediate product, is the fraction of parent compound which appears as the

intermediate product, and k2

 is the rate parameter for intermediate product degradation. These equations can be

solved to yield

D D k P

k k

e

k P

k k

e

kt kt m m = + −

F

H

G I

K

J − −

− −

0

1 0

1 2

1 0

1 2

2 1 α α ρ ρ

(46)

where D0

 is the initial concentration of the intermediate product.

For most situations involving chlorinated compounds, k2

 is substantially less than k1. For the time domain where

net degradation of the intermediate product is occurring, the first term on the right side of Equation 46 dominates,

therefore the concentration is approximated as

D D k P

k k

e

k t m ≈ + −

F

H

G I

K

J −

0

1 0

1 2

2 α ρ

(47)

Making the same substitutions used in deriving Equation 42 yields

W

A

F nu

k

D

D

k P

D

k k

≈ + −

F

H

G

G

I

K

J

J 2

0 1

0

1 2

ln

α

(48)

As TCE degrades, DCE and vinyl chloride often appear in solution at concentrations corresponding to a few

percent of the TCE originally present (Focht et al., 1996). Depending on the actual amount of conversion (α) and

the values of the other parameters, Equation 48 may indicate the need for more reactive material than would be

suggested by Equation 42. For example, with no intermediate products initially present and with k2

/k1 = 0.2,

α= 0.03, and a remediation goal of P0

/D = 1,000, Equation 48 will indicate a weight-per-area that is 2.6 times the

amount calculated with Equation 42.

Evaluation of Existing Iron Permeable Barriers Systems

The iron PRBs installed to date range in weight-per-area from 19 to 850 lb/ft2 (Table 10), in good agreement with

the calculations of the previous section.

Table 10. Summary Information on Full-Scale Permeable Reactive Barriers Using Fe(0).

Site Iron

(tons)

Capture Area

(ft2)

W/A

(lb/ft2)

Coffeyville, KS 70 7500 19

Belfast, N. Ireland 15 650 46

Denver, CO 580 11,700 99

Elizabeth City, NC 300 2900 210

Mountainview, CA 90 660 270

Lowry AFB, CO 45 270 330

Upstate NY 45 270 380

Moffett AFB, CA 96 475 400

Somersworth, NH 65 250 520

Sunnyvale, CA 220 520 850

C -
81

Each of these W/A values can be understood in site-specific terms. Those sites with low W/A are examples of sites

with low ground-water velocity and without significant concerns regarding slow-reacting intermediate products.

The Sunnyvale, California, site has a high W/A value for several reasons. First, the reaction rate parameter (k) is

low because the principal contaminants, DCE and VC, degrade much slower than TCE. Second, the ground-water

velocity is relatively high, at approximately 0.8 ft/day. The safety factor applied in this case was 4 (Warner et al.,

1995).

Cost Estimation

Characterizing a plume in terms of the weight-per-area of reactive material required lends itself to evaluating costs

on a per-area basis. Installation costs for impermeable barriers are often expressed in this manner. For example,

shallow impermeable barriers are often quoted to cost between $10/ft2

 and $25/ft2 excluding mobilization. In this

section, it is shown how this practice may be extended to permeable reactive barriers.

Many of the iron PRBs installed thus far are of the funnel-and-gate design, in which one or more discrete

permeable trenches (“gates”) is installed along with impermeable containment walls (“funnels”) to direct ground

water through the gate. The funnel-and-gate cost analysis presented here will focus on three elements: the iron

itself, the creation of the gate, and the funnel. All incremental cost components will be expressed on a per-area

basis. It is important to understand that these costs are on the basis of the total cross-sectional area of the plume

being treated, not the area of the gate.

The cost of reactive material on a per-area basis is simply the product of its cost per weight and the weight-per-area

(W/A) required:

Reactive materials

cost per plume

cross sectional

area

−

F

H

G

G

G

G

I

K

J

J

J

J

= ×F

H IK Cost per weight W

A a f (49)

A typical price for the granular iron currently being used for PRBs is $375/ton. This translates into incremental

costs of $9/ft2

 for a W/A of 50 lb/ft2, $47/ft2 for a W/A of 250 lb/ft2, and $188/ft2 for a W/A of 1,000 lb/ft2.

The cost of installing a gate is most readily quoted on a per-volume basis. To obtain the gate cost on the basis of

plume cross-sectional area, this value is multiplied by the required gate volume per plume area. The gate volume

required per plume area is simply W/A divided by the bulk density of granular iron. So,

Gate installation

cost per plume

cross sectional area

Installation cost

per volume −

F

H

G

G

I

K

J

J

=F

H

G I

K

J ×

W A

ρb (50)

where ρb

 is the bulk density of granular iron, typically about 160 lb/ft3. In most funnel-and-gate systems installed

thus far, gate installation has proved expensive. By using a trench box method and confined entry procedures for the

final stages of installation, costs of $1,000/yd3

 have not been unusual. This corresponds to a cost of $12/ft2 to install

50 lb/ft2

 of iron gate, $58/ft2 to install 250 lb/ft2 and $231/ft2 to install 1,000 lb/ft2.

Tables 11 and 12 present cost estimates for several scenarios. In each table, costs are estimated for three values of

W/A: 50, 250, and 1,000 lb/ft2

. For Table 11, relatively high installation costs are used: $25/ft2 for funnels and

$1,000/yd3

 for gates. For Table 12, significantly lower costs were assumed: $10/ft2 for funnels and $200/yd3 for

gates.

Comparing Tables 11 and 12 demonstrates that there is considerable incentive for employing less costly means of

installing both the impermeable and permeable components of funnel-and-gate systems. Further cost savings may

be realized in some cases by designing continuous PRBs, thereby eliminating the gate installation cost altogether,

assuming that the costs of installing continuous permeable barriers are similar to the costs of installing impermeable

barriers (funnels).

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Table 12. Cost Elements of Funnel and Gate Systems: Low Estimate (Gate @ $200/yd3; Funnel @ $10/ft2)

Table 11. Cost Elements of Funnel and Gate Systems: High Estimate (Gate @ $1000/yd3; Funnel @ $15/ft2)

Incremental Costs

($/ft2)

Component

W/A =

50 lb/ft2

W/A =

250 lb/ft2

W/A =

1000 lb/ft2

Iron 9 47 188

Install gate 12 58 231

Funnel 25 25 25

Sum 46 130 444

Incremental Costs

($/ft2)

Component

W/A =

50 lb/ft2

W/A =

250 lb/ft2

W/A =

1000 lb/ft2

Iron 9 47 188

Install gate 2 12 46

Funnel 10 10 10

Sum 21 69 244

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83

Appendix D. Acronyms

AFB Air Force Base

AI&M Ada Iron & Metal

ASF Anderson-Schulz-Flory distribution

bgs below ground surface

Ci initial contaminant concentration of the influent solution

C0 contaminant concentration in solution at time t = 0

CF chloroform (CHCl3

)

CM chloromethane (CH3

Cl)

CT carbon tetrachloride (CCl4

)

DCE dichloroethene (C2

H2Cl2)

tDCE trans-dichloroethene (C2

H2Cl2)

cDCE cis-dichloroethene (C2

H2Cl2)

DCM dichloromethane (CH2

Cl2)

DI deionized

DO dissolved oxygen

DOD Department of Defense

DOE Department of Energy

DNAPLs dense nonaqueous phase liquids

DL detection limit

e- electron

ECC electrochemical corrosion cell

EDX Energy Dispersive X-ray Analysis

f

oc fraction of organic carbon

F&G funnel-and-gate

FAME fatty acid methyl ester analyses

GAC granular activated carbon

GC/MS gas chromatography/mass spectrometry

GPR ground-penetrating radar

GW ground water

h hour

HC hydrocarbon

HCB hexachlorobenzene

HPLC high-performance liquid chromatography

ISTZ in situ treatment zone

ITRC Interstate Technology and Regulatory Cooperation Workgroup

k rate constant

kobs observed rate constant

kSA surface-area-specific rate parameter

MBS Master Builders Supply iron

MCL maximum concentration limit

MW monitoring well

NACE National Association of Chemical Engineers

NAS National Academy of Sciences

ND nondetectable

NERL National Exposure Research Laboratory

NMR Nuclear Magnetic Resonance Spectroscopy

NRMRL National Risk Management Research Laboratory

O&M operations and maintenance

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84

ORD Office of Research and Development

ORNL Oak Ridge National Laboratory

OSHA Occupational Safety and Health Administration

OSWER Office of Solid Waste and Emergency Response (U.S. EPA)

PAC powdered activated carbon

PCE tetrachloroethene (perchloroethene)

PEI-silica polyethylenimine-coated silica beads

PRB permeable reactive barrier

psi per square inch

PV pore volume

QAPP Quality Assurance Project Plan

r

f retardation factor

ROD Record of Decision

rpm revolutions per minute

RSF reactant sand-fracturing

RT residence or retention time

RTDF Remediation Technologies Development Forum

RWQCB Regional Water Quality Control Board (State of California)

rxn reaction

RZF reaction zone formation technologies

SA surface area

SBR stirred batch reactor

S.H.E standard hydrogen electrode

SEM scanning electron microscopy

SI saturation index or indices

SIMS secondary ion mass spectroscopy

SITE Superfund Innovative Technologies Evaluation

SMAD cryoparticles metal vapor–solvent codeposition method (in preparation of active metal surfaces)

SMVs or SVOCs semivolatile organics

soln solution

SPM scanning probe microscopy

t

1/2 half life

TCA trichloroethane

TCE trichloroethene

TCM trichloromethane

TCP trichloropropane

TDS total dissolved solids

TER Technology Evaluation Report

TIO Technology Innovation Office (U.S. EPA)

TLC thin layer chromatography

TOC total organic carbon

UMTRA Uranium Mill Tailings Remedial Action Program

US ultrasound

USCG United States Coast Guard

vol or v volume

VC vinyl chloride

VISITT Vendor Information System for Innovative Treatment Technologies

VOCs volatile organic compounds

VOAs volatile organic aromatics

W/A weight per area

XRD X-Ray Diffraction

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85

XPS X-Ray Photoelectron Spectroscopy

y year

ZHE zero head-space extractors

zpc point of zero charge

ZVI zero-valent iron

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86

Appendix E. Glossary

anode The electrode in an electrochemical cell toward which anions are drawn and where

oxidation occurs.

Arrhenius equation An equation that expresses the logarithmic relationship between the rate constant of a

reaction and the reciprocal of the temperature (expressed in K).

bentonite The term bentonite is used as a commercial name for clays that are predominantly sodium

montmorillonite. Wyoming bentonite is the most common drilling fluid additive used in

the water well industry.

biotite An important mineral of many intrusive igneous rocks, pegmatites, lamprophyres, some

lavas and metamorphic rocks.

calcite CaCO3. A sedimentary mineral formed by chemical precipitation.

Specific Gravity ≈ 2.95 g/cm3

.

cell potential The voltage of an electrochemical cell.

cation A positively charged ion, attracted toward the cathode in an electrolytic cell.

cathode The electrode in an electrochemical cell toward which cations are drawn and where

reduction occurs.

Eh or Nernst equation, expresses the relationship between the standard redox potential of a

given redox couple, its observed potential, and the concentration ratio of its electron-donor

and electron-acceptor species.

electrochemical cell A system containing an oxidation-reduction reaction in which oxidation and reduction

reactions are physically separated and the transferred electrons pass through an electrical

circuit.

Fenton’s Reagent A way to generate OH• by Fenton reaction:

Fe2+

aq + H2

O2 → Fe3+

aq + OH• + OH- (1)

Irradiation with light λ<580nm effects photoreduction of Fe3+ to Fe2+ together with the

production of OH• radicals

Fe3+

aq+ H2

O + h → OH• + Fe2+

aq + H+ (2)

goethite An iron mineral of the general formula FeOOH (alpha-FeOOH, see lepidocrocite).

Specific Gravity 4.28 g/cm3

.

half-life The time required for 50% of a material or compound to undergo transformation or decay.

hematite An iron oxide mineral, Fe2

O3, corresponding to an iron content of approximately 70%.

Specific Gravity ≈ 5.26 g/cm3

.

iron metal Variously designated as Fe0

, Fe(0), or zero-valent iron. The most common reactive media

in the majority of field scale and commercial PRB implementations. Sources of Fe used in

experiments and installations referenced in this document are: Ada Iron and Metal,

Aldrich, Alfa Aesar, Peerless, Fluka, Fisher, VWR, MBS.

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87

Kobs This is the rate at which the contaminants are observed to degrade.

kSA Normalized kobs to Fe surface concentration yields a specific rate constant, kSA, for a

number of solvents.

Lewis acid A species that accepts a pair of electrons to form a covalent bond.

Lewis acid-base theory The idea that an acid is a species that accepts a pair of electrons to form a covalent bond

and a base is a species that donates a pair of electrons to form a covalent bond.

lepidocrocite gamma-FeOOH (see goethite)

MINTEQA2 A geochemical equilibrium speciation model for dilute aqueous systems. It is an update of

MINTEQ, that was developed by combining the fundamental mathematical structure of

MINEQL with the thermodynamic data base of WATEQ3. For more information see

http://www.cee.odu.edu/cee/model/minteq.html

montmorillonite An aluminosilicate clay mineral in the smectite group

muscovite [KAl2

[AlSi3O10](OH,F)2]. One of the more common minerals in rocks, especially plutonic

igneous rocks rich in silica or aluminum and low or medium to high grade metamorphic

rocks. Specific Gravity≈2.8 g/cm3

.

Nernst equation An equation that expresses the exact electromotive force of a cell in terms of the activities

of products and reactants of the cell.

oxidation (1) a reaction in which there is an increase in valence resulting from a loss of electrons.

Contrast with reduction. (2) A corrosion reaction in which the corroded metal forms an

oxide; usually applied to reaction with a gas containing elemental oxygen, such as air.

passivation (1) A reduction of the anodic reaction rate of an electrode involved in corrosion. (2) The

process in metal corrosion by which metals become passive. (3) The changing of a

chemically active surface of a metal to a much less reactive state.

PHREEQC A program for aqueous geochemical calculations. For more information go to

http://water.usgs.gov/software/phreeqc.html

potential-pH diagram A plot of the redox potential of a system versus the pH of the system, compiled using

thermodynamic data and the Nernst equation. The diagram shows regions within which the

metal or mineral itself or some of its compounds are stable.

pyrite An iron sulfide mineral with the general formula of (FeS2

). Specific Gravity≈5.0 g/cm3.

Common in plutonic, volcanic, sedimentary and metamorphic rocks.

redox potential (Eh) The subscript H is used to emphasize that the potential only has meaning in reference to the

standard hydrogen electrode reaction.

saturation indices Gauges the potential for minerals to precipitate. A negative SI indicates undersaturation

with respect to the particular mineral phase, while a positive SI indicates oversaturation.

siderite An iron oxide mineral with the general formula FeCO3. Specific Gravity≈3.95 g/cm3

.

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88

surfactant A surface-active agent; usually an organic compound whose molecules contain a hydrophilic

group at one end and a lipophilic group at the other.

vermiculite A mica-like silicate mineral of the general formula (Mg,Fe2+,Al)3

(Si,Al)4O10(OH)2

zeolites Complex inorganic framework mineral structures made up from SiO4 and AlO4 tetrahedra

joined together to form a series of interconnected channels and pores. Small molecules can

diffuse through the zeolite and undergo chemical reactions catalyzed by active sites in the

channel walls. The high acidity of their protonated form, coupled with the heterogeneous

and controllable nature of the reactions which proceed within them has made them ideal

choices for a wide range of catalytic processes.

E -
89

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