vvEPA
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In Situ Treatment of
Soil and Groundwater
Contaminated with Chromium
Technical Resource Guide
In
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EPA 625/R-00/004
October 2000
IN SITU TREATMENT OF SOIL
AND GROUNDWATER
CONTAMINATED WITH CHROMIUM
TECHNICAL RESOURCE GUIDE
Center for Environmental Research Information
National Risk Management Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
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Notice
The U.S. Environmental Protection Agency through its Office of Research and
Development funded and managed the research described here under Contract
No. 68-C7-0011, Work Assignment 27, to Science Applications International
Corporation (SAIC). It 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.
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Foreward
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 this mandate, 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 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.
This publication has been produced as part of the Laboratory's strategic long-term
research plan. It is published and made available by EPA's Office of Research and
Development to assist the user community and to link researchers with their clients.
E. Timothy Oppelt, Director
National Risk Management Research Laboratory
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Abstract
Chromium is the second most common metal found at sites for which Records of
Decision have been signed. At many industrial and waste disposal locations,
chromium has been released to the environment via leakage and poor storage
during manufacturing or improper disposal practices. Industrial applications most
commonly use chromium in the hexavalent chromium [Cr(VI)] form, which is acutely
toxic and very mobile in groundwater. Groundwater extraction and treatment has
traditionally been used to remediate chromium-contaminant plumes. This method,
while providing interception and hydraulic containment of the plume, may require
long-term application to meet Cr(VI) remediation goals and may not be effective at
remediating source-zone Cr(VI).
New information and treatment approaches have been developed for chromium-
contaminated soil and groundwater treatment. The purpose of this report is to bring
together the most current information pertaining to the science of chromium
contamination and the in situ treatment and control of sites with groundwater
and/or soil contaminated with chromium. A number of available in situ technologies
or treatment approaches use chemical reduction and fixation for chromium
remediation. These include geochemical fixation, permeable reactive barriers
(PRBs), and reactive zones. Other types of in situ treatment that are under
development include enhanced extraction, electrokinetics, biological processes that
can be used within PRBs and reactive zones, natural attenuation, and
phytoremediation.
Detailed discussions of these in situ technologies are contained in the report. Each
discussion consists of a technology description with its advantages and
disadvantages, status, and performance and cost data. A comparative summary
of the status of the technologies is presented in Table 3-1. More conventional ex
situ approaches and other proven and well-documented technologies for chromium
treatment or control are not reviewed within this report. The emphasis in this report
is on innovative in situ approaches for chromium remediation that are not as well
documented, but have been demonstrated or are being developed.
It should be noted that this report is not a design document, but a resource guide.
Although it does contain design and cost information, it is primarily intended to
enable concerned parties, regulators, scientists, and engineers to evaluate the
potential use of various treatment technologies, or combinations of these
technologies, to clean up chromium-contaminated sites effectively.
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Table of Contents
Section Page
Notice i
Foreward ii
Abstract iii
Figures vi
Tables vii
Acronyms, Abbreviations, and Symbols viii
Acknowledgments xi
1 - INTRODUCTION
1.1 Purpose 1
1.2 Background and Regulatory Overview 1
1.3 Scope of the Guide 3
2-CHROMIUM IN THE ENVIRONMENT 4
2.1 Sources and Extent of Contamination 4
2.2 Chromium Chemistry 5
2.3 Chromium Treatment and Remediation Approaches 11
2.4 Site Characterization Requirements 14
3 -TECHNOLOGIES FOR IN SITU TREATMENT 17
3.1 Geochemical Fixation 20
3.1.1 Technology Description 20
3.1.2 Status 23
3.1.3 Performance and Cost Data 23
3.2 Permeable Reactive Subsurface Barriers (Treatment Walls) 30
3.2.1 Technology Description 30
3.2.2 Status 36
3.2.3 Performance and Cost Data 38
3.3 Reactive Zones 44
3.3.1 Technology Description 44
3.3.2 Status 52
3.3.3 Performance and Cost Data 52
3.4 Soil Flushing/Chromium Extraction 56
3.4.1 Technology Description 56
3.4.2 Status 59
3.4.3 Performance and Cost Data 59
3.5 Electrokinetics 61
3.5.1 Technology Description 61
3.5.2 Status 64
3.5.3 Performance and Cost Data 65
3.6 Natural Attenuation 66
IV
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TABLE OF CONTENTS (continued)
Section Page
3.6.1 Technology Description 66
3.6.2 Status 69
3.6.3 Performance and Cost Data 69
3.7 Phytoremediation 70
3.7.1 Technology Description 70
3.7.2 Status 70
3.7.3 Performance and Cost Data 70
4-REFERENCES 71
Appendix Page
A - SOURCES OF ADDITIONAL INFORMATION 77
B - TECHNOLOGY AND VENDOR CONTACT INFORMATION 82
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LIST OF FIGURES
2-1 Metals Most Commonly Present in all Matrices at Superfund Sites 4
2-2 Eh-pH Diagram for Chromium 6
2-3 The Chromium Cycle in the Environment 7
2-4 Chromium Reduction and Fixation 11
2-5 Concentration Versus Pumping Duration or Volume Showing Tailing
and Rebound Effects (Cohen et al., 1994) 12
2-6 Conceptual Geochemical Model of Zones in a Contaminant Plume 13
3-1 Schematic of In Situ Chromium Remediation 21
3-2 Plot of Hexavalent Chromium Contaminant Plume at the Valley Wood
Preserving Site - January, 1998 25
3-3 Plot of Hexavalent Chromium Contaminant Plume at the Valley Wood
Preserving Site - November, 1999 26
3-4a Plume Capture by a Funnel and-Gate System. Sheet Piling Funnels
Direct the Plume Through the Reactive Gate 32
3-4b Plume Capture by a Continuous Trench System. The Plume Moves
Unimpeded Through the Reactive Gate 32
3-5 Breakdown of Inorganic Contaminants Addressed by PRBs 37
3-6 Breakdown of Types of PRB Projects Addressing Chromium
Remediation 37
3-7 ISRM Treatment System Diagram 42
3-8 In Situ Reactive Zones Curtain Design Concept 46
3-9 Gravity Feed of Reagents When the Contamination is Shallow 49
3-10 Multiple Cluster Injection Points When Contamination is Deep 49
3-11 Schematic of Molasses-based Injection System at Central Pennsylvania Site 54
3-12 Plot of Hexavalent Chromium Contaminant Plume at Central Pennsylvania
Site - January, 1997 55
3-13 Plot of Hexavalent Chromium Contaminant Plume at Central Pennsylvania
Site - July, 1998 55
3-14 Schematic of In Situ Flushing System 57
3-15 Electrokinetic Remediation Process 62
VI
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LIST OF TABLES
Table
1-1 Cleanup Goals (Actual and Potential) for Total and Leachable Metals ... 2
2-1 Cation Exchange Capacities for Soils - Components and Types 9
2-2 Recommended Analytical Methods 15
3-1 Status of In Situ Technologies for Treatment of Chromium Contamination 18
3-2 Total Chromium Concentrations (mg/L) of Paired-Column Effluent as a
Function of Throughput 24
3-3 Comparison of Granular vs. Foam Iron Reactive Media for PRBs 35
3-4 Impacts of Various Geologic/Hydrogeologic Parameters on the Design
of an In Situ Reactive Zone 48
3-5 Results of Cr(VI) Extraction Studies by Surfactants and Hydrotropes from
Elizabeth City Soil 60
3-6 Cr(VI) Extraction From Columns by Water, Surfactants Alone, and
Surfactant Solubilized DPC 61
3-7 Case Studies from Gil 65
VII
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Acronyms, Abbreviations, and Symbols
(aq)
(s)
AEC
BaCrO4
BOAT
Ca
CaCO3
CCA
Cd
CEC
CERCLA
CH4
Cl
cm
CMC
CN
Co
CO2
Cr
Cr(lll)
Cr(VI)
Cr20/2
Cr04-2
CrOHx
Cu
Dl
DMRB
DOC
DOE
DPC
Eh
EIX
EPA
Fe
Fe(ll)
Fe(lll)
Fe°
ft
ft2
ft3
g
H2
HCrO4
HSO3-
ICP/AES
ICP/MS
aqueous
solid
less than
greater than
anion exchange capacity
barium chromate
Best Demonstrated Available Technology
calcium
calcium carbonate
copper chromium arsenate
cadmium
cation exchange capacity
Comprehensive Environmental Response, Compensation, and Liability Act
methane
chloride
centimeter
critical micelle concentration
cyanide
cobalt
carbon dioxide
chromium
trivalent chromium (reduced form)
hexavalent chromium (oxidized form)
dichromate
chromate
chromium hydroxide (numerous species)
copper
deionized
dissimilatory metal-reducing bacteria
dissolved organic carbon
U.S. Department of Energy
diphenyl carbazide
redox potential
electrochemical ion-exchange
U.S. Environmental Protection Agency
iron
ferrous iron (reduced form)
ferric iron (oxidized form)
zero-valent iron
foot
square foot
cubic foot
grams
hydrogen
hydrochromate
hydrosulfite
Inductively coupled plasma/atomic emission spectrometry
Inductively coupled plasma/mass spectrometry
VIM
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Acronyms, Abbreviations, and Symbols (Continued)
in
ISEE
ISRM
kg/m3
kW-hr/m3
kWh
L/min
Ibs/ft3
M
m
m2/ft3
m2/g
m2/$
MCL
meg
mg/kg
mg/L
mg/h
m/W
mmol/L
Mn
MnO
MnO2
mV
N2
Ni
NO2-1
N03-1
NPL
O&M
OH
OSC
Pb
PBAT
PH
PLC
P04-3
ppb
PRB
RCRA
ROD
ROOM
RPM
RTDF
S
s2o6-2
SDL
Se
SITE
inch
In Situ Electrokinetic Extraction
In Situ Redox Manipulation
kilogram per cubic meter
kilowatt-hour per cubic meter
kilowatt-hour
liter per minute
pounds per cubic foot
molar
meter
square meters per cubic foot
square meters per gram
square meters per dollar
maximum contaminant level
milli equivalent
milligram per kilogram
milligram per liter
milligram per hour
millimolar
millimole per liter
manganese
manganese oxide
manganese dioxide
milli-volt
nitrogen
nickel
nitrite
nitrate
National Priorities List
Operation and Maintenance
hydroxide
On-Scene Coordinator
lead
Permeable Barriers Action Team
negative Iog10 of hydrogen ion concentration
programmable logic controller
phosphate
parts per billion
permeable reactive barrier
Resource Conservation and Recovery Act
Record of Decision
organic acid
Remedial Project Manager
Remediation Technologies Development Forum
sulfur
metabisulfite
Sandia National Laboratories
selenium
Superfund Innovative Technology Evaluation
IX
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Acronyms, Abbreviations, and Symbols (Continued)
SMZ surfactant-modified zeolite
SO4"2 sulfate
TCE trichloroethene
TCLP Toxicity Characteristic Leaching Procedure
TOC total organic carbon
pg/L microgram per liter
USCG U.S. Coast Guard
WET Waste Extraction Test
yd3 cubic yard
Zn zinc
ZVI zero-valent iron
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Acknowledgments
This Technical Resource Guide was prepared under the direction and coordination of Mr. Douglas
Grosse of the U.S. Environmental Protection Agency (EPA) National Risk Management Research
Laboratory (NRMRL) in Cincinnati, Ohio. Mr. Grosse served as the project officer. Contributors to
and reviewers of this Guide were Mr. Grosse; Ms. Marta Richards of NRMRL; Dr. Robert Puls of
NRMRL, Ada, Oklahoma; Mr. Ralph Howard, EPA Region IV RPM; Dr. Andy Davis of Geomega,
Inc.; Mr. Ishwar Murarka of Ish, Inc.; and Mr. Jim Rouse of Montgomery Watson. Ms. Jean Dye of
NRMRL provided final editorial review, and Mr. John McCready of NRMRL designed the cover.
This Guide was prepared for the EPA NRMRL by Mr. Kyle Cook, Mr. Robert Sims, Ms. Allison
Marten, and Mr. John Pacetti of Science Applications International Corporation (SAIC). Kyle Cook
was the SAIC Work Assignment Manager. Special thanks is given to Ms. Debbie Seibel, Ms.
Christina Hudson, Ms. Elizabeth Wunderlich, and Mr. Richard Dzija of SAIC for their administrative
and technical assistance in preparing this Guide.
XI
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Section 1
INTRODUCTION
1.1 Purpose
Contamination of soil and groundwater by
chromium is a significant problem in the United
States. Conventional groundwater treatment
approaches such as pumping and ex situ
treatment have been used to address this
problem. The poor performance of pump-and-
treat systems in the mid-1980s provided the
driving force for research of subsurface
processes in order to develop more efficient
groundwater remediation strategies and
techniques. As a result, new information and
treatment approaches have been developed
for chromium-contaminated soil and
groundwater treatment. However, much of this
information is either scattered in the literature
or not directly available, making it difficult for
decision makers to access. The purpose of this
document, therefore, is to bring together the
most current information pertaining to the
science of chromium contamination and the
treatment and control of sites with groundwater
and/or soil contaminated with it. It is hoped that
this information will enable remedial project
managers (RPMs); on-scene coordinators
(OSCs); state, local, and Indian tribal
regulators; technology vendors; consultants;
private organizations; and citizens to evaluate
the potential use of treatment technologies to
clean up chromium-contaminated sites
effectively.
1.2 Background and Regulatory
Overview
Groundwater can become contaminated with
metals directly by infiltration of leachate from
land disposal of solid wastes, sewage or
sewage sludge; leachate from mining wastes;
seepage from industrial lagoons; and spills and
leaks from industrial metal processing or wood
preserving facilities. Numerous waste and site
conditions control and influence the leachability
of the metals and wastes and their transport
into groundwater (Evanko and Dzombak, 1997).
Several categories of technologies exist for the
remediation of metals-contaminated soil and
water. These include: isolation, immobilization,
toxicity reduction, physical separation, and
extraction. Combinations of one or more of
these approaches are often used for more
efficient and cost-effective treatment of a
contaminated site. In situ treatment methods for
metals-contaminated soil and groundwater are
being tested and will be applied with increasing
frequency (Evanko and Dzombak, 1997). While
a great deal of progress has been made, a
number of needs and issues still need to be
addressed before in situ soil and groundwater
remediation technologies will be most effective.
For those technologies that have proven
successful in field pilot-scale demonstrations,
an important next step is to take the technology
to a larger scale demonstration and to use it for
an actual site remediation. For those
technologies that have been implemented at
one or a few sites, the next goal is to make
them more commonly accepted practices for
site remediation. An important part of this
transition is the acquisition of credible cost data
and the development and evaluation of cost
models that can be used for system design
(Schmelling, 1999).
Soil and groundwater cleanup goals for
chromium (Cr) and other metals can be defined
as total and leachable metals. Actual and
potential soil cleanup goals for chromium are
presented in Table 1-1. Total metals goals are
levels that can be applied directly to
contaminated soil. There is some variability in
these goals for total chromium. The total metals
analysis determines the level of metals
contamination expressed as mg metal/kg soil.
This analysis does not take into consideration
1
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the leachability of the metals contaminants.
For leachable metals, the leachate levels are
applied to solid wastes such as soil after
extracting and testing the material with the
specified analytical methods. Leachable metals
may be determined by different analytical
procedures, as shown in Table 1-1, that
measure the concentration of metals expressed
in mg/L of the leachable material. The
maximum contaminant level (MCL) and
Superfund Site Goals are applied to
groundwater.The observed variation in cleanup
goals has at least two implications in regard to
technology alternative evaluation and selection.
First is the importance of identifying the target
metals contaminant state (leachable vs. total
metal), the specific type of test and conditions,
and the numerical cleanup goals early in the
remedy evaluation and process. Second, the
effectiveness of a technology for meeting total
or leachable treatment goals should be viewed
with some caution in light of the degree in
variation in goals and the many factors that
affect mobility of the metals (USEPA, 1997).
Table 1-1. Cleanup Goals (Actual and Potential) for Total and Leachable Metals
Description
Total Metals Goals (mg/kg)
Background (Mean)
100
Background (Range)
1 to 1000
Superfund Site Goals
6.7 to 375.0
Theoretical Minimum Total Metals to Ensure TCLPa Leachate
threshold (i.e., TCLPx 20)
100
California Total Threshold Limit Concentration
500
Leachable Metals (ug/L)
TCLPa Threshold for RCRA Waste (SW 846, Method 1311)
5000
Extraction Procedure Toxicity Test (EP Tox) (Method 1310)
5000
Synthetic Precipitate LeachateProcedure (Method 1312)
Multiple Extraction Procedure (Method 1320)
California Soluble Threshold Leachate Concentration
5000
MCLb
100
Superfund Site Goals
50
a TCLP = Toxicity Characteristic Leaching Procedure
b The maximum permissible level of contaminant in water delivered to
any user of a public system.
No specified level and no example cases identified.
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In uncontaminated natural waters, the range of
chromium concentrations is quite large. The
median value in unpolluted freshwater or
seawater is usually low, less than 50 mmol/L.
For most natural waters, the chromium
concentration is below the 50 ^g/L value
(approximately 1 mmol/L) recommended for
drinking water by the Commission of European
Communities, the World Health Organization,
or the U.S. Environmental Protection Agency
(Richard and Bourg, 1991). As shown in Table
1-1, the Superfund Site Goal for groundwater is
50^g/L, and the maximum contaminant level is
100/vg/L
1.3 Scope of the Guide
Section 2 of this Guide provides more
background information on the sources of
chromium contamination and the extent of the
problem in the United States.
Section 2 also presents a detailed examination
of chromium chemistry in order to help the
reader better understand the behavior of
chromium in the environment, its
characterization, and its relationship to
treatment approaches.
Section 3 explores the latest in situ
technological approaches for dealing with
chromium site remediation, the emphasis of
this Guide. Finally, additional sources of
information and studies are listed in
Appendices A and B for those interested in
acquiring more information.
This document is not intended to be a
comprehensive description of in situ
technological approaches for dealing with
chromium site remediation. Rather, it is
intended to be used as a resource guide in
conjunction with other references, such as
those listed in Appendix A, the opinions of
technology experts, and site-specific
information. Therefore, the reader is cautioned
that information provided in this document,
including costs, is specific to the study cited,
and may not be directly transferable to other
applications.
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Section 2
CHROMIUM IN THE ENVIRONMENT
2.1 Sources and Extent of
Contamination
Chromium is an important industrial metal used
in diverse products and processes (Nriagu,
1988). At many industrial and waste disposal
locations, chromium has been released to the
environment via leakage and poor storage
during manufacturing or improper disposal
practices (Palmer and Wittbrodt, 1991; Calder,
1988). The first instances of groundwater
contamination were associated with chromium
plating operations at aircraft manufacturing
facilities during World War II. Many of these
plumes have moved thousands of feet over the
past 50 years, with little or no natural
attenuation (Rouse, 1997). The National
Priority List (NPL) of 1986 developed by the
U.S. Environmental Protection Agency (EPA)
presents approximately 1,000 sites in the U. S.
that pose significant environmental health
risks (Allen etal., 1995).
About 40 percent of these sites have been
reported to have metals problems. The majority
of these reported metals are combined with
organics, but a significant number reported are
metals only or metals with inorganics. The
metals most often cited as problems are lead,
chromium, arsenic, and cadmium. According to
EPA data for sites for which Records of
Decision (RODs) have been signed, chromium
is the second most common metal found at
these contaminated sites (U.S. EPA, 1996).
Figure 2-1 summarizes the occurance and
distribution of metals at these sites. Currently,
the principal sources of chromium-
contaminated soils and groundwater are
electroplating, textile manufacturing, leather
tanning, pigment manufacturing, wood
preserving, and chromium waste disposal
(U.S. EPA, 1997).
500-
450-
400-
350-
300-
250 -
200-
150-
100-
50-
0
460
306
235 226
224
201
154
Lead Chromium Arsenic Zinc Cadmium Copper Mercury
Contaminants
From U.S. EPA, 1996
Figure 2-1 Metals most commonly present in all matrices at Superfund sites.
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The use of metals in waterborne wood
preserving solutions has increased over time,
with consumption in 1995 exceeding all other
processes combined (AWPI, 1996). By far the
most widely used wood preserving formulation
is copper chromium arsenate (CCA). Due to
operating procedures that were standard
practices at the time, nearly all wood preserving
plants 20 years or older present some degree
of soil and groundwater contamination. At least
71 wood preserving sites have been listed on
the Comprehensive Environmental Response,
Compensation, and Liability Act (CERCLA)
NPL (Federal Register, 1996). At least 678
additional sites exist at which wood preserving
operations have been or are currently being
conducted; contamination may be present at
many of these sites as well (U.S. EPA, 1997c).
The most significant groundwater and soil
contamination problem associated with the use
of CCA is hexavalent chromium [Cr(VI)] which
is acutely toxic, mutagenic, and carcinogenic in
the environment. It is also very soluble, mobile,
and moves at a rate essentially the same as
the groundwater (Palmer and Puls, 1994). In
contrast, the reduced form of Cr(VI), trivalent
chromium [Cr(l 11)], has relatively low toxicity and
is immobile under moderately alkaline to slightly
acidic conditions. Industrial applications most
commonly use chromium in the Cr(VI) form,
which can introduce high concentrations of
oxidized chromium (chromate) into the
environment. Cr(VI) does not always readily
reduce to Cr(ll I) and can exist over an extended
period of time.
2.2 Chromium Chemistry
This section describes the basic chemistry
involved with the various oxidation states of
chromium to account for the behavior of this
metal in the natural environment, and links this
information to in situ technologies discussed in
the following sections of this Guide.
Aqueous Chemistry and pH Effect
Chromium has a unique geochemical behavior
in natural water systems. Cr(lll) is the most
common form of naturally occurring chromium,
but is largely immobile in the environment, with
natural waters having only traces of chromium
unless the pH is extremely low. Under strong
oxidizing conditions, chromium is present in the
Cr(VI) state and persists in anionic form as
chromate. Natural chromates are rare.
However, the use of Cr(VI) in wood preserving
CCA solutions, metal plating facilities, paint
manufacturing, leather tanning, and other
industrial applications has the potential to
introduce high concentrations of oxidized
chromium to the environment (Rouse and
Pyrih, 1990; Palmer and Wittbrodt, 1991).
Redox potential Eh-pH diagrams present
equilibrium data and indicate the oxidation
states and chemical forms of the chemical
substances which exist within specified Eh and
pH ranges. Figure 2-2 is an Eh-pH diagram for
chromium. The data presented in Figure 2-2
are derived from parameters representing
typical aqueous conditions. Although the
diagram implies that the boundary separating
one species from another is distinct, the
transformation is so clear cut. Concentration,
pressure, temperature, and the absence or
presence of other aqueous ions can all affect
which chromium species will exist. A measure
of caution must be exercised when using this
diagram as site-specific conditions can
significantly alter actual Eh-pH boundaries.
-------
Ill
1.2 -
1 "
0.8
0.6 -
0.4
0.2
| j Hexavafent
\ Trivalenl Chromium
-0.8
0248
and
Figure 2-2 Eh-pH diagram for chromium.
Palmer and Wittbrodt (1991) claim that
chromium exists in several oxidation states
ranging from 0 to 6. Under reducing conditions,
Cr(lll) is the most thermodynamically stable
oxidation state. However, Cr(VI) can remain
stable for significant periods of time. In soils
and aquifer systems, the most prevalent forms
are the trivalent and hexavalent oxidation
states.
Cr(lll) exists in wide Eh and pH ranges. Palmer
and Wittbrodt (1991) have determined that the
following Cr(lll) species exist with respect to
pH. Cr(lll) predominates as ionic (i.e., Cr+3) at
pH values less than 3.0. At pH values above
3.5, hydrolysis of Cr(l 11) in a Cr(l I l)-water system
yields trivalent chromium hydroxy species
[CrOhT2, Cr(OH)2+, Cr(OH)3°, and Cr(OH)4-].
Cr(OH)3° is the only solid species, existing as
an amorphous precipitate. The existence of the
Cr(OH)3 ° species as the primary precipitated
product in the process of reducing Cr(VI) to
Cr(lll) is paramount to the viability of in situ
treatment using reactive zone technology, such
as microbial bioreduction. Cr(lll) can form
stable, soluble (and thus mobile), organic
complexes with low to moderate molecular
weight organic acids (i.e., citric and fulvic
acids). The significance of these complexes is
that they allow Cr(l 11) to remain in solution at pH
levels above which Cr(lll) would be expected to
precipitate (Bartlett and Kimble, 1976a; and
James and Bartlett, 1983a).
-------
Reactions and Mechanisms in Aquifer
Systems
The chemistry of aqueous chromium in an
aquifer is complicated, interactive between soil
and water, and cyclic in the reactions that occur
as they relate to solid and dissolved phases
and the various oxidation states present. The
"Chromium Cycle" is presented in Figure 2-3.
Understanding this chemical process is
important in the decision-making process in
determining which treatment technology (either
singly or in combination) to use.
The two major oxidation states of chromium
which occur in the environment are Cr(lll) and
Cr(VI). According to Bartlett (1991), the
following conditions exist: "Cr(VI) is the most
oxidized, mobile, reactive, and toxic chromium
state. In general, under non-polluting
conditions, only small concentrations of Cr(VI)
species exist [the result of oxidation of natural
Cr(lll)], with Cr(lll) species being the most
prevalent forms. Moist soils and sediments in
partial equilibrium with atmospheric oxygen
contain the conditions needed in which
oxidation and reduction can occur
simultaneously. Cr(lll) species may be oxidized
to Cr(VI) by oxidizing compounds that exist in
the soil (i.e., manganese dioxide - MnO2), while
at the same time Cr(VI) species may be
reduced to Cr(lll) by MnO2 in the presence of
reduced manganese oxide (MnO) and organic
acids from soil organic matter." In addition, the
reduction of Cr(VI) to Cr(lll) in soils will most
likely occur as a result of reduction by soil
organic matter (including humic acid, fulvic
acid, and humin), soluble ferrous iron [Fe(ll)],
and reduced sulfur compounds. Therefore, it is
important to understand the geochemical
environment of any site where Cr(VI) is likely to
occur.
cr(iii)
Soil Adsoprtion,
§
Cr(VI)'
Uptake
in Bbsphere
5J (Anton with PcJ, SOf }
(Anton Exchange
SsuKe: from larfett, 199
Figure 2-3 The Chromium Cycle in the environment.
-------
The success of geochemical fixation treatment
techniques is based on forming insoluble non-
reactive chemical species. Precipitation and
adsorption result in fixation or solid-phase
formation of Cr(lll), each depending on the
physical and chemical conditions existing in the
aquifer system. Precipitation reactions can be
further divided into three types: pure solids
such as Cr(OH)3° (amorphous precipitation);
mixed solids or coprecipitates such as
CrxFe1.x(OH)3; and high molecular weight
organic acid complexes such as humic acid
polymer (Palmer and Wittbrodt, 1991 and
James and Bartlett, 1983b). Pure solid Cr(lll)
hydroxide precipitates result from changes in
the Eh-pH parameters (Figure 2-2).
Chromium hydroxide solid solutions may
precipitate as coprecipitates with other metals,
rather than pure Cr(OH)3°. This is especially
true if oxidized iron [Fe(lll)] is present in the
aquifer; it will generate an amorphous
hydroxide coprecipitate in the CrxFe1.x(OH)3
form (Palmer and Wittbrodt, 1991). This
chemical reaction is particularly important due
to the potential for Fe(ll) to be oxidized to the
ferric state as previously discussed. Fe(ll) is the
most common oxidation state of dissolved iron
in natural subsurface waters as well as aquifer
minerals. Advantage is taken of this chemical
reaction when employing permeable reactive
barrier (PRB) in situ treatment of ground water.
Zero-valent iron (Fe°) metal is used to reduce
Cr(VI) to Cr(lll) and complex the Cr(lll) as a
Fe(lll) hydroxide coprecipitate.
Insoluble organic acid complex precipitates with
Cr(lll) and soil humic acid polymers are
generally quite stable and present a barrier to
Cr(lll) oxidation to Cr(VI). Cr(lll) is tightly bound
and immobilized by insoluble humic acid
polymers.
The name given to this complexation process is
chrome tanning because chromium has
replaced aluminum in the tanning of leather.
The chrome tanning of soil organic matter limits
the tendency for Cr(lll) to become oxidized and
for the organic matter to be decomposed (Ross
etal., 1981).
Adsorption reactions generally consist of cation
exchange capacity (CEC) mechanisms for
Cr(lll) species and anion exchange capacity
(AEC) mechanisms for Cr(VI) species.
Adsorption generally involves cation exchange
of Cr(lll) as Cr+3 or hydroxy ionic species onto
hydrated iron and manganese oxides located
on the surface of clay soil particles. In CEC
mechanisms, an aquifer mineral lattice or
hydrated iron and manganese oxides located
on the surfaces of fine-grained soil particles
adsorb cations. Competition with other similar
ions is possible and may limit the absorption of
one particular species. Understanding CEC
mechanisms is critical when considering in situ
treatment technologies, such as soil
flushing/chromium extraction and electrokinetic
remediation. Generally, the lower the CEC of
the soil, the better suited the soil for
remediation by these technologies. Table 2-1
presents the CECs for various soil
classifications (Dragun, 1988). The soil organic
matter component of soil provides the greatest
CEC, followed by the clay minerals vermiculite,
saponite and montmorillinite. Clay offers the
greatest CEC of all the soil types.
8
-------
Table 2-1. CECs for Soils - Components and Types.
CEC
(meq/1 OOg )
Soil Clays
Chlorite
Illite
Kaolinite
Montmorillonite
Oxides and Oxyhydroxides
Saponite
Vermiculite
10-40
10-40
3-15
80-150
2-6
80-120
100-150
Soil Type
Soil Organic Matter
Sand
Sandy Loam
Loam
Silt Loam
Clay Loam
Clay
>200
2-7
2-18
8-22
9-27
4-32
5-60
In addition to soil cation exchange mechanisms
for Cr(lll) species adsorption, soil anion
exchange is possible for adsorption of Cr(VI)
anions [i.e., hydrochromate (HcrO4~) and
chromate (CrO4~2)]. These species exchange
with chloride (Cl"), nitrate (NO3"), sulfate (SO/),
and phosphate (PO4~3). Griffin et al. (1977)
studied the effect of pH on the adsorption of
Cr(VI) by the clay minerals kaolinite and
montmorillonite, and found adsorption was
highly pH dependent; the adsorption of Cr(VI)
decreased as pH increased, and the
predominant Cr(VI) species adsorbed was
HCrO4". Bartlett and Kimble (1976b) also found
that while chromate is tightly bound compared
with anions such as Cl" or NO3~, it can be
released by reaction of the soil with PO4~3. The
presence of orthophosphate prevented the
adsorption of Cr(VI) anions, presumably by
competition for the adsorption sites. They
concluded that the behavior of Cr(VI)
remaining in soils is similar to that of
orthophosphate, but unlike phosphate, Cr(VI) is
quickly reduced by soil organic matter, thus
becoming immobilized. Cr(VI), they state, will
remain mobile only if its concentration exceeds
both the adsorbing and the reducing capacities
of the soil.
Sulfate adsorption on kaolinite also varied with
pH, although not as strongly as for chromate.
Zachara et al. (1988) suggested that, although
2 -2
SO4" and CrO4 compete for adsorption sites
on noncrystalline iron oxyhydroxide, SO4 and
CrO4 bind to different sites on kaolinite and,
thus, do not compete for the same site. Studies
by Zachara et al. (1989) of the adsorption of
chromate on soils found the following:
Chromate adsorption
decreasing pH.
increased with
-------
• Soils that contained higher concentrations
of aluminum and iron oxides showed
greater adsorption of Cr(VI).
• Chromate binding was depressed in the
presence of dissolved SO4~2 and inorganic
carbon, which compete for adsorption sites.
The importance of dissolved SO4~2 as an
inhibitor to chromate binding becomes apparent
in Section 3. The technology described in
Section 3.1.1, geochemical fixation, uses
sodium metabisulfite as a reductant to reduce
Cr(VI) in groundwater to Cr(lll). Additional
sulfur-based reductants are described in
Section 3.1.3. Also, one of the Permeable
Reactive Barrier (PRB) technologies described
in Sections 3.2.1 and 3.2.3 uses sodium
dithionite to reduce Fe(lll) to Fe(ll), which in
turn reduces Cr(VI) to Cr(lll). In all these
processes, SO4~2 is produced as a byproduct
and can act as an inhibitor to Cr(VI) adsorption,
thus allowing any residual Cr(VI) to remain in
the mobile phase to be reduced to Cr(lll).
In situ treatment methods for chromium-
contaminated soil and groundwater generally
involve the reduction of Cr(VI) to Cr(lll) with
subsequent fixation of Cr(lll). Figure 2-4
presents examples of natural and chemical-
induced reduction of Cr(VI) to Cr(lll) and the
mechanisms of subsequent fixation of Cr(lll).
The permanence of fixation must be evaluated
since Cr(lll) [as low molecular weight organic
acid complexes (i.e., chromium citrate)] can
migrate to the surface and reoxidize to Cr(VI) in
the presence of manganese dioxide.
Manganese dioxide (MnO2) forms naturally in
the upper vadose zone by reduced manganese
oxide (MnO) reacting with atmospheric oxygen.
Bartlett (1991) states "the marvel of the
chromium cycle in soil is that oxidation and
reduction can take place at the same time."
This is an important principle for the application
of in situ technologies for the treatment
(reduction) of Cr(VI) and permanent fixation of
Cr(lll).
Figure 2-3 illustrates the apparent paradox of
simultaneous oxidation and reduction of
chromium. As shown, Mn(IV) (as MnO2)
oxidizes Cr(lll) to Cr(VI). However, under
normal dry soil conditions, mobile Cr(lll) [i.e.,
Cr+3 or chromium citrate] will not oxidize to
Cr(VI) in the presence of MnO2. Mobile Cr(lll)
will not oxidize to Cr(VI) in the presence of
MnO2 unless the soil is moist and the MnO2
surface present in the soil is fresh (i.e.,
amorphous rather than crystalline form)
(Bartlett, 1991). Additionally, Mn(lll)-organic
acid complexes reduce Cr(VI) to its trivalent
form. Mn(lll) is formed when Mn(ll) reacts with
Mn(IV) in the presence of organic acids formed
from soil organic matter (Bartlett, 1991). The
cycle repeats itself as the Cr(lll) formed may be
chelated by low molecular weight organic acid
complexes (e.g., citric acid) and thus, be mobile
enough to migrate to the soil surface and
consequently oxidize to Cr(VI).
Bartlett (1991) states that as long as all Cr(VI)
has been reduced and all Cr(lll) is bound by
decay-resistant organic polymers, the
chromium will remain inert and immobile,
provided that oxygen is excluded. In other
words, sealing of a landfill on the bottom to
prevent leaching of chromium is unnecessary
as long as the top is sealed.
10
-------
Natifal
Reduction"
Chemical
Seduction
CrfVIj
In Environment
Mn(ill)
Seid
Figure 2-4 Chromium reduction and fixation.
2.3 Chromium Treatment and
Remediation Approaches
Palmer and Wittbrodt (1991) discussed several
remediation techniques for chromium-
contaminated sites. Applicable to many sites is
a pump-and-treat method. The technology
works by extracting contaminated groundwater,
usually over long time periods, and providing
hydraulic control (containment) of a
contaminant plume. Initially, the concentration
of the contaminant is high in the effluent, but
with continued pumping, the concentration
decreases significantly. These residual
concentrations remain above the MCLs, and
can persist for long periods of time, called
"tailing." This same phenomenon was observed
by Stollenwerk and Grove (1985) in their
laboratory and batch column experiments.
Figure 2-5 shows tailing and rebound effects
during and after groundwater pumping. Tailing
is the result of several physical and chemical
processes:
Differential time for contaminants to be
advected from the boundary of the plume to an
extraction well: Groundwater flows, not only in
response to an extraction well, but also to the
natural hydraulic gradient. As a result, not all of
the water in the vicinity of an extraction well
enters the well. There is a limited area, the
capture zone, from which the water is
captured, and a stagnation point, located
downgradient from the well, where the velocity
toward the well equals the velocity induced by
the natural gradient. The net velocity is zero,
and there is little change in the concentration
of the contaminant during the pump-and-treat
remediation. In addition, the groundwater
velocity of a volume of water moving from the
11
-------
Apparent Residua
Contaminant
Coneentfrntton
Figure 2-5 Concentration versus pumping duration or volume showing tailing and
rebound effects (Cohen et al., 1994).
edge of the plume to the extraction well is
greater than a volume of water traveling along
a streamline on the outside of the capture
zone. The time it takes the contaminated water
to flow is controlled by the thickness of the
aquifer, the rate of groundwater extraction, the
natural groundwater gradient, and the gradient
induced or impacted by other
injection/extraction wells.
Diffusive mass transport within spatially
variable sediments: Geologic materials are
typically heterogeneous; groundwater moves
through higher permeable layers while water in
lower permeable layers remains immobile.
Contaminants that have remained in the
subsurface for extended periods of time
migrate to the lower permeable layers by
molecular diffusion. During pump-and-treat,
clean water is moved through the more
permeable layers at a relatively high rate, while
removal of the contaminants from the lower
permeable lenses is limited by the rate of
diffusion into the higher permeable layers; thus
maintaining the concentration of the
contaminant, often above the established MCL.
Mass transfer from residual solid phases in the
aquifer: Contaminants can exist in the
subsurface in relatively large reserves as solid
phase precipitates. A likely reserve for
chromium contaminated sites is barium
chromate (BaCrO4), the source of the barium
either coming from contamination or from the
natural soil.
Palmer and Wittbrodt (1991) conducted a
study at a United Chrome Products site and
suggested that the Cr(VI)-contaminated
groundwater was in equilibrium with BaCrO4.
Column leaching tests of the contaminated
soils showed a significant leveling of the Cr(VI)
concentrations, indicating that a solid phase
may be controlling the concentration in the
extraction water.
Sorption/desorption processes: As discussed
previously, Cr(VI) exists in solution as the
anions HCrO4~, CrO4~2, and dichromate
(Cr2Cy2), and is adsorbed onto the soil matrix.
As the concentration of Cr(VI) decreases, it
becomes more difficult to remove the Cr(VI).
12
-------
The use of in situ technologies such as
chemical enhancement of the pump-and-treat
method (the addition of reductant or extracting
agent) may be desirable to overcome the
tailing phenomenon and reduce the overall
time required for remediation. However, the
cause of tailing at a given site needs to be
determined and quantified. For example, if the
tailing is controlled by physical processes such
as differential travel time along streamlines, or
heterogeneity of the soil, then chemical
enhancement may not be advantageous.
Further, regulatory agencies may require the
removal of the chemical enhancer. This is
especially true if the chemical enhancer or its
byproducts exceed the concentration(s)
applicable water quality standards.
of
Typically, chromium-contaminated sites consist
of three zones: (1) source zone soils where the
concentrated waste resides; (2) the
concentrated portion of the groundwater
plume; and (3) the diluted portion of the
groundwater plume (Sabatini et al.,
1997).Figure 2-6 illustrates these three zones
of contamination. Applying conventional pump-
and-treat remediation methods to all three
regions would be highly inefficient. An
integrated technology approach would
probably be best suited for full-scale site
remediation.
Derived from: Rouse et al., 1996 and Sabatini et al., 1997.
Figure 2-6 Conceptual geochemical model of zones in a contaminant plume.
13
-------
Available Technologies
A number of in situ technologies or approaches
use chemical reduction/fixation for chromium
remediation. These include geochemical
fixation, PRBs, reactive zones, and natural
attenuation. Understanding the Chromium
Cycle presented in Section 2.2 and site
characteristics presented in Section 2.4 is
critical for the use of these approaches,
especially natural attenuation. Chemical
reduction/fixation remediation techniques do
not remove chromium from the aquifer system,
but are designed to immobilize chromium
precipitates by fixing them onto aquifer solids
or reactive media, thereby reducing chromium
in groundwater. Other types of in situ treatment
that are available or under development for
remediation of chromium-contaminated sites
include soil flushing/enhanced extraction,
electrokinetics, and biological processes
including phytoremediation. Biological
processes include bioreduction,
bioaccumulation, biomineralization, and
bioprecipitation which use specific substrates to
drive the treatment and effect the reduction,
uptake, or precipitation of Cr(VI) based on the
principles presented in Section 2.2. These
processes can be utilized within PRBs and
reactive zones. Phytoremediation utilizes plant
uptake of chromium contamination as the in
situ treatment approach.
A detailed discussion of these in situ
technologies is presented in Section 3 of this
Guide. More conventional ex situ approaches
are not reviewed in this Guide.
2.4 Site Characterization
Requirements
The remediation site should be characterized to
determine how suitable it is for Cr(l II) fixation or
for other treatment applications. Chemical
characterization should include the following:
Site Characterization: Site characterization
should include a determination of total organic
carbon (TOC) and dissolved organic carbon
(DOC) in groundwater and soil. Both tests will
indicate not only the availability of soluble
organic ligands for Cr(lll) complexing, which
provides a mobilization vehicle for potential
oxidation to Cr(VI), but also the availability of
more complex organic matter which has the
potential for reduction of Cr(VI) to Cr(lll). The
particulate (or solid fraction) of organic carbon
in the aquifer can be determined by subtracting
DOC from TOC. A total Cr(VI) reducing
capacity of the soil should be determined to
measure the portion of organic matter in the
soil that is oxidizable by Cr(VI). The Cr(VI) not
reduced is titrated with Fe(ll).CEC should be
measured to determine if sites are available for
the Cr(lll)-hydroxy cation complexes to adsorb
onto the soil particles. Other tests that can be
performed as needed are porosity, grain size,
soil moisture, and total manganese.
Groundwater: Both contaminated and treated
groundwater should be analyzed for total
chromium and Cr(VI); Cr(lll) is determined by
subtracting the results of the Cr(VI) from the
total chromium values. Eh and pH should also
be determined.
So/7: Like the groundwater, both contaminated
and treated aquifer solids and unsaturated soil
should be analyzed for total and Cr(VI).
Additional tests should be conducted for pH
and the amount of dissolved Cr(lll) that is
mobile (not fixed). Further, in order to
determine if, and how much of, the Cr(VI) was
reduced, a mass balance should be performed.
Other soil tests that can be performed as
needed are the standard chromium oxidation
test; Cr(lll) oxidizable by excess MnO2; and
oxidizability of inert Cr(lll). The methods for
these tests, along with their rationale, are
presented in Bartlett (1991).
Table 2-2 lists recommended analytical
methods. The tests presented in Table 2-2 are
meant to be a guide as to the types of
analytical parameter measurements that are
helpful in understanding the ongoing
geochemical processes at a site as they relate
chromium remediation. The list of analytical
parameters is not meant to be comprehensive,
but should provide a good foundation.
14
-------
Table 2-2. Recommended Analytical Methods
Sample Type
Site Characterization
Groundwater
Pre- and Post-Treatment
Soil
Pre- and Post-Treatment
Soil
Post-Treatment (Leachate)
Analyte
TOC (water)
TOC (soil)
DOC (water)
Particulate Organic Carbon
Soil pH
Groundwater pH
CEC
Total Cr(VI) reducing capacity
by soil
Total manganese (soil)
Total Chromium
Cr(VI)
Cr(lll)
Total Chromium
Cr(VI)
Trivalent Chromium
Available Cr(lll) (to be
mobilized)
Cr(lll)
EPA415.1 or415.2
SW-846 modified 9060
0.45 urn filter, then EPA 41 5. 1 or 41 5.2
TOC minus DOC
SW-846 9045C (use distilled water)
EPA 150.1
EPA 9081
Walkley-Black method
Digest: SW-846 3050B, 3051, or 3052
Analysis: SW-846 7460, 601 OB, or 6020
0.45 urn filter, Digest: SW-846 3020A
Analysis: SW-846 71 91
-or-
Digest: SW-846 3005A
Analysis: SW-846 601 OB or 6020
0.45 urn filter
Analysis: SW-846 71 96A
Total Cr-Cr(VI)
Digest: SW-846 3050B, 3051, or 3052
Analysis: SW-846 7090, 601 OB, or 6020
Digest: SW-846 3060A
Analysis: SW-846 71 96A
Total Cr-Cr(VI)
Prep: K2H - citrate extract (Bartlett, 1991)
Analysis: SW-846 71 96A
Leachate: Title 22 Waste Extraction Test
(WET)
Digest: SW-846, 301 OA
Analysis: SW-846 7090, 601 OB, or 6020
-or-
Digest:SW-846 3020A
Analysis: SW-846 71 91
15
-------
There may be other measurements that could
be added to this list to obtain a better
understanding of the geochemical processes at
any one site. Whenever possible, EPA SW-846
analytical methods were proposed because
their performance effectiveness has been
validated, the methods are commonly used in
the literature and provide a high degree of data
comparability, and they have been approved by
EPA. For some analytical parameters there is
more than one proposed analytical method.
This provides greater flexibility to the users of
this Guide. For example, there are two choices
for digestion techniques and three choices for
analysis for total chromium in groundwater.
Although all the digestion and analytical
techniques are valid, one may choose to
analyze for chromium at a site at which low
levels are expected by SW-846, 7191 (atomic
adsorption, furnace) rather than 601 OB or 6020
inductively coupled plasma/atomic emission
spectrometry or mass spectrometry (ICP/AES
or ICP/MS), because the detection limit is lower
by 7191. The digestion method to use for
analytical method 7191 is SW-846 3020A, and
digestion method SW-846 3005A is to be used
with analytical methods 601 OB and 6020. In
addition to site chemistry, it is also critically
important for in situ technology implementation
to understand the contaminant distribution and
geologic setting. This includes geologic
structure, stratigraphy, and ground-water
hydrogeology. Complicated geology and low
permeability zones will influence how a
technology is applied and its treatment
effectiveness. Laboratory and pilot-scale tests
can help to determine the effectiveness of the
treatment on the contaminated matrix prior to
full-scale application of the technology.
16
-------
Section 3
TECHNOLOGIES FOR IN SITU TREATMENT
This section discusses innovative technologies
that can potentially be used for in situ
remediation of chromium-contaminated soil and
groundwater. More conventional technologies
for treatment or control of chromium and metals
contamination such as pump-and-treat,
containment, solidification/stabilization, and
even thermal treatments, are not reviewed here
because they are either ex situ approaches or
are proven and well-documented elsewhere.
The emphasis in this Guide is on innovative in
situ approaches that are not as well
documented in order to provide the reader with
the latest information for decision-making
purposes.
For clarification, individual technologies and
two technology groups are presented in this
section. The individual technologies are
geochemical fixation, soil flushing/chromium
extraction, electrokinetics, natural attenuation,
and phytoremediation. The technology groups
are subsurface PRBs and reactive zones.
Numerous biotic and abiotic applications can
be utilized within these technology groups and
are discussed. The emphasis in this Guide is
on the technologies or treatment approaches
proven or demonstrated for in situ chromium
remediation - geochemical fixation and PRBs.
These technologies have been utilized at full-
scale and are generally supported by
performance and cost data.
Reactive zones, soil flushing, and
electrokinetics are developing technologies that
have been demonstrated at pilot-scale and
have limited performance and cost data
available. Natural attenuation and
phytoremediation are emerging technologies
under study for in situ chromium remediation.
All are covered briefly in this Guide. Each
technology or technology group is presented in
a separate section that provides a technology
description, advantages and limitations of using
the technology, its status, and performance and
cost information, if available.
No technology may be able to remove 100
percent of the contaminants that are present at
a site. Consequently, it is important to
determine the benefits of partial mass removal
and relate this to risk reduction. Many sites will
require different approaches to different parts
of the site. As a result, there is a need to better
understand how to link technologies together to
achieve site cleanup in the most cost-effective
manner (Schmelling, 1999).
Table 3-1 summarizes the status of the
technologies discussed in this section.
17
-------
Table 3-1. Status of In Situ Technologies for Treatment of Chromium Contamination
Proven Technologies
GEOCHEMICAL FIXATION
PRBs
• Chemical Reduction &
Fixation (Reactive Media)
• Chemical Reduction &
Fixation (ISRM)
• Adsorption & Chemical
Reduction (Zeolite/ZVI)
Developing Technologies
REACTIVE ZONES
• Chemical Reduction &
Fixation (Reactive agent)
• Bioreduction (Molasses)
• Bioaccumulation (Yeast)
Bioprecipitation (DMRB)
• Biomineralization
SOIL FLUSHING/
EXTRACTION
ELECTROKINETICS
TREATMENT
ZONE1
S, C, D
C, D
C, D
C, D
S, C, D
S, C, D
C
S, C
S, C
S, C, D
S, C
CONTAMINATED
MEDIA
Groundwater
Soil
Groundwater
Groundwater
Groundwater
Groundwater
Groundwater
Groundwater
Groundwater
Groundwater, Soil
Groundwater
Soil
Groundwater
Soil
STATUS 2
DE, FSA, FTS
DE, FTS
DE, FSA, FTS
PE, FTS
PE, BTS
DE, FSA, FTS
DE, FSA, FTS
PE, BTS
PE, BTS
PE, FTS
DE, FSA; PE, BTS
DE, FSA
DE, FTS, BTS
PE, FTS
TREATMENT COST
ESTIMATE3
$4/m3 ($3/yd3) for
saturated zone using
ferrous sulfate4
- O&M 70-90% less
than P&T per year.
-$3 mil./10yr vs. $9
mil. for P&T.
-NA
NA
$400,000/3 yr vs.
$4 mil/20 yr for P&T.
NA
NA
NA
$60to$170/ton, or
$83 to $237/m3 at
assumed soil density
of 100lb/ft35
$25 to $300/m3, or
$19to$229/yd35
REGULATORY ACCEPTANCE
Gaining acceptance: depends
on site evaluation results and
reductants proposed for use.
Physical treatment walls are
starting to gain regulatory
acceptance. R&D supported by
EPA and other organizations.
May be difficult to get regulatory
approval without further
research and demonstration of
these treatment techniques.
Some acceptance for water-only
flushing applications. More
difficult for surfactants or other
additives. Hydraulic control is an
issue.
Case-by-case acceptance for a
developing technology. Needs
more research and
demonstration.
00
-------
Table 3-1. (continued)
Emerging Technologies
NATURAL ATTENUATION
PHYTOREMEDIATION
TREATMENT
ZONE1
C, D
C, D
CONTAMINATED
MEDIA
Groundwater
Soil
Groundwater
Soil
STATUS 2
PE, FTS
PE
PE, FTS
PE, FTS
TREATMENT COST
ESTIMATE3
Detailed site
characterization and
performance
monitoring costs
$10to$60/ton, or
$18to$104/m3at
assumed soil density
of 100lb/ft35
REGULATORY ACCEPTANCE
Beginning to gain regulatory
acceptance in general.
Uncertain for Cr remediation -
highly dependent on site
characteristics.
May be difficult to get regulatory
approval without further
research. Early stage of
development for metals
remediation.
CD
1 S=Source zone soil or groundwater, C=Concentrated groundwater plume, D=Diluted groundwater plume
2DE=Demonstrated Effectiveness, PE=Potential Effectiveness, FSA=Full-scale Application, FTS=Field-scale Treatability Study, BTS=Bench-scale Treatability Study
3These costs were compiled from studies presented in this document and are for informational purposes only; actual treatment costs are highly site and application specific and may vary
considerably from costs presented here. Costs in this document have not been adjusted to year 2000 dollars and may not include profit.
4 Does not include treatability and design costs, which were significant.
5 These cost estimates may not include indirect costs such as permits, treatment of residues, and site preparation.
NA=Not Available
-------
3.1 Geochemical Fixation
3.1.1 Technology Description
The goal of this technology is to reduce Cr(VI)
in groundwater and contaminated soil to the
more thermodynamically stable Cr(lll). The
reduced chromium is expected to
geochemically fix onto aquifer solids. The
technology is based upon the concept of
extracting contaminated groundwater and
treating it above ground, followed by reinjection
of the treated groundwater into the aquifer. The
reinjected groundwater is dosed with reductant
to reduce any residual Cr(VI) contamination
remaining in the interstitial water. This
technique, if successful, yields Cr(VI)
concentrations in groundwater below that of
drinking water standards (50 ug/L) at a fraction
of the aquifer pore volume throughput required
by typical pump-and-treat methods.
The success of the in situ chromium
geochemical fixation technology depends on
the ability of the applied reductant to reduce
Cr(VI) in groundwater to Cr(lll), and on the
capacity of the reduced chromium to fix onto
the aquifer solids. The total chromium
concentration in the aquifer system is not
decreased, but chromium is precipitated and
fixed (immobilized) onto aquifer solids as Cr(lll)
so that it is not available in the groundwater.
Some information is available indicating how
sodium metabisulfite might act as a reductant in
the subsurface. In general, sulfur compounds
such as sulfide and sulfite reduce Cr(VI). For
sulfides to reduce Cr(VI), Fe(ll) must be present
to act as a catalyst. Thus, in aquifer systems
where iron sulfides are present, reduction of
Cr(VI) may occur. However, because the rate of
reaction is slow, the iron sulfide reduction
process may not effectively treat large volumes
of water.
According to Palmer and Wittbrodt (1991), the
following reactions occur. In the presence of
excess sulfite, the reduction of Cr(VI) follows
the reaction:
(1) 6H+ + 2HCrO4- + 3HSCV (excess) -*•
2Cr+3 + 2SO4-2 + S2O6-2 + 6H2O
The metabisulfite (S2O6~2) formed by the above
reaction can then reduce oxidized Fe(lll) to
Fe(ll), if it is present. This situation allows for
potential reduction of Cr(VI) by Fe(ll), as
previously described. In the presence of excess
Cr(VI), the reduction to Cr(lll) by sulfite follows
the reaction:
(2) 5H+ + 2HCrO4- (excess) + 3HSCV --
2Cr+3
3SO4-2
5H2O
Therefore, the process of using sodium
metabisulfite should reduce the Cr(VI) to Cr(lll)
in situ, provided there are sufficient iron and
manganese oxide adsorption sites within the
aquifer treatment zone to which the Cr(lll) can
affix.
Using the technology, groundwater is extracted,
treated with a chemical reagent, and reinjected
along the contaminant plume perimeter (see
Figure 3-1). As the treated water is directed
towards the center of the plume, Cr(VI) is
reduced to Cr(lll), its less soluble form. The
zone of contamination is driven inward by the
reaction front, leaving behind an increasingly
larger clean water zone. Alternatively, injection
can occur in the high concentration areas
(source zones) to effect a more rapid
remediation (Brown et al., 1998). In situ
geochemical fixation can be applied to source
or core zones, the concentrated or active zone,
or the dilute or neutralized zone of the
contaminant plume.
20
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Reductant Treated Water
Advancing Front of
Heducton! Treawd Water
Figure 3-1 Schematic of in situ chromium remediation.
The technology has also been used for in situ
reduction and fixation of Cr(VI) retained as
dissolved chromium in interstitial water in the
vadose (unsaturated) zone. In such
applications, groundwater is recovered from
near the area of contaminated soil, amended
with the selected reductant, and percolated
through the soil. The percolation displaces a
portion of the moisture from the vadose zone
(hence the need to recover the groundwater
from near the area of contamination), but also
reduces and fixes the residual chromium onto
the soil as Cr(lll). This technique has been
applied in geologic material as diverse as silt,
glacial outwash sand, fractured siltstone, and
limestone with extensive solution features. The
effectiveness of the approach is best measured
by the use of pressure/vacuum lysimeters to
sample the interstitial fluids (Rouse, 1999).
Geochemical properties that should be
evaluated for use of the technology include
acid and base neutralizing potential, iron and
manganese hydrous-oxide content, TOC, and
CEC (Rouse and Pyrih, 1990). Analysis of the
anions sulfate, phosphate, chloride, and nitrate
and AEC may also be valuable for the
evaluation of their competition with Cr(VI) for
sorption cites on aquifer solids, and how well
Cr(VI) can be mobilized for treatment
(reduction). The site hydrogeology must also be
well understood; successful application
requires a good knowledge of the hydrogeology
and expertise at designing and physically
implementing the process. Section 2 of this
Guide provides more detail about process
geochemistry and site characterization needs.
21
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Extraction and injection wells are required to
remove contaminated groundwater for
aboveground treatment and to reinject the
treated and amended water into the treatment
zone. Extraction wells from an existing pump-
and-treat system can often be utilized. The
aboveground system consists of the selected
treatment apparatus for initial chromium
removal, tanks for storing treatment chemicals
and extracted and treated groundwater, and a
reductant dosing system for the treated water.
A sheltered area or building is needed to
perform the on-site chemical tests, handle and
prepare the chemical reductant, and to store
spare parts such as backup pumps and valves.
A significant advantage to using in situ
geochemical fixation is that it has the potential
to substantially reduce (up to 75 percent) the
time required to remediate chromium-
contaminated sites to meet cleanup goals, and
thus reduce treatment (operating) costs. Other
advantages are listed below.
Advantages
•D Better hydraulic control is achieved by the
reinjection of treated water around the
contaminated plume, forming a "ridge" of
treated water.
•D Since water is reinjected, the same
gradient can be established at lower
pumping rates, thereby avoiding
"stranding" of chromium in dewatered
portions of the previously saturated
aquifer.
•D In situ reduction of residual Cr(VI) in
interstitial void spaces can be achieved.
•D The amount of treatment plant sludge for
disposal is reduced since more Cr(VI) is
reduced and precipitated in situ.
•D The surface discharge of treated water is
reduced since water is reinjected (Rouse,
1994; Rouse and Pyrih, 1990; Rouse,
Leahy and Brown, 1996).
Limitations
•D Aquifer materials must have the ability to
permanently "fix" Cr(lll).
•D Reduced Cr(lll) could re-oxidize to Cr(VI)
under certain conditions (presence of
manganese dioxide [MnO2]); however, this
has not been observed in the field.
•D Need regulatory approval for reinjection of
pumped and treated groundwater.
•D Aquifer material heterogeneities
(stratification, etc.) makes design and
treatment more difficult.
•D Aquifer solids must be porous to water
flow.
•D Ferrous sulfate-based reductants may
result in iron precipitation and clogging of
aquifer pore spaces.
•D Excess reductant or reductant byproduct
may have to be removed if undesirable or
if it exceeds groundwater MCLs.
22
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3.1.2 Status
This technology has undergone testing and is
currently being employed at contaminated sites
using calcium/sodium polysulfide and sodium
metabisulfite chemical reductants to effect the
conversion of Cr(VI) to Cr(l 11) (Rouse, 1997). In
situ remedial systems are or have been in
operation at chromium contaminated sites in
California, Indiana, Maryland, Maine and South
Australia, and are planned for Michigan and
additional sites in California.
The technology is also being tested and
applied on soil and groundwaterata Superfund
site in South Carolina using infiltration and well
injection methods (Howard, 1998). Ferrous
sulfate is the reductant of choice for this
application. The technology will become part of
the final remedial design for the site.
In situ geochemical fixation can also be applied
to a site remediation using more passive
techniques such as in situ PRBs, or reactive
zones. These approaches are presented in
Sections 3.2 and 3.3 of this Guide.
3.1.3 Performance and Cost Data
Delaware River
At the site of a former paper mill on the
Delaware River, the Cr(VI) concentration in the
perched aquifer was 85 mg/L (85,000 fjg/L).
After treatment by reduction and precipitation
using ferrous sulfate, Cr(VI) levels across most
of the site were reduced to 50 pg/L (below the
drinking water standard). These levels have
been maintained for over 4 years (Brown et. al.,
1998). This was the first commercial application
of the iron reduction process for treating Cr(VI)
in soils and groundwater. The total cost of the
project was $250,000, which is approximately
would have been required to initiate pump-and-
equal to the cost of capital equipment that
treat. For perched groundwater treatment, the
application of an acidified ferrous sulfate
heptahydrate was carried out using a
combination of infiltration galleries, addition
point/wells, and a vertical trellis network. No
groundwater was extracted for treatment and
reinjection; therefore, no treatment sludge was
generated that required disposal.
Indiana
A wood treatment plant site in Indiana had four
areas of soil and groundwater contaminated
with Cr(VI), a by-product of the CCA solution
used to treat wood. Four plumes from the site
threatened the domestic drinking wells. With
conventional pump-and-treat technologies,
cleaning up this site could have taken more
than 10 years at a cost of several million
dollars. However, the site plumes were
remediated using geochemical fixation by
extracting and mixing groundwater with a
reductant and reinjecting the treated water
through upgradient soils and groundwater. In
the critical "off-site" plume, Cr(VI) was reduced
to below the residual health-based groundwater
criterion of 0.1 mg/L in 2 months using a
calcium polysulfide reductant. Smaller on-site
plumes were treated by addition of sodium
bisulfide reductant via injection wells. After 3
months of treatment, chromium levels in all
three wells had gone from concentrations of up
to 0.80 mg/L to less than 0.01 mg/L. Hot-spots
were also being treated. It should be noted that
in some locations, Cr(VI) levels were already at
or near the groundwater criterion, so significant
reductions were not required. Work on these
projects began in 1995 and was completed in
1997 at a total cost of approximately $600,000
(Rouse etal., 1999).
23
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Central California
As much as 6,000 pounds of Cr(VI) are
estimated to have been released to the site soil
and groundwaterfrom wood treating operations
at the Valley Wood Preserving site in Turlock,
California. A groundwater pump-and-treat
system maintained hydraulic control of the
plume, treated nearly 9 million gallons of water,
and recovered about half (3,000 pounds) of the
Cr(VI). It was estimated that recovery of the
remaining contamination would have required
another 10 years of treatment (Brown et al.,
1998). Bench-scale parallel column tests were
conducted to evaluate the efficacy of the use of
a reductant on chromium contamination. Three
paired column tests were performed, each
using actual chromium-contaminated aquifer
material. For each paired column, one column
was flushed with demineralized water to
simulate the typical response of a pump-and-
treat or "clean water sweep" remedial
approach, and the other with a mild reductant
solution designed to achieve in situ reduction
and fixation of Cr(VI).
Table 3-2 presents the results from the column
tests. The data demonstrate that the reductant-
treated water achieved effluent concentrations
less than the drinking water standard of 0.10
mg/L after 4 to 6 pore volumes (a pore volume
is the volume of groundwater in the pore
spaces of a defined aquifer zone), while
columns flushed with only water exceeded the
standard after 14 to 17 pore volumes (Rouse,
1994). Next, the efficacy and potential of in situ
geochemical fixation as a treatment alternative
was demonstrated by conducting an on-site
pilot test. The on-site pilot test consisted of a
"push-pull" test. This test involves removing a
measured quantity of water from a
contaminated well, treating it with the
appropriate reductant (predetermined),
returning it to the well, and after a period of
time, collecting samples from the well to
measure the effect of the added chemical
reductant. Based on these results, a sulfur-
based reductant (sodium metabisulfite) is being
used to remediate Cr(VI) at this site (Brown et
al., 1998).
Table 3-2. Total Chromium Concentrations (mg/L) of Paired-
Column Effluent as a Function of Throughput
Pore
Volume
1
2
3
6
9
10
14
17
Soil A
Demineralized
Water
270
65
38
19
15
-
-
-
Reductant
Solution
210
17
6.8
0.05
-
0.06
-
-
SoilC
Demineralized
Water
2,800
220
18
10.5
-
1.35
-
0.22
Reductant
Solution
2,800
120
1.05
0.01
-
0.02
-
-
SoilD
Demineralized
Water
1,700
120
23
4.1
-
0.49
0.19
-
Reductant
Solution
1,700
34
0.62
0.02
-
<0.01
-
-
- = No sample taken.
24
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A full-scale treatment system has been in
operation since January 1998 and has reduced
the maximum chromium concentrations by
more than an order of magnitude. In situ
treatment from February 1998 through October
1999 has resulted in a reduction in plume size
and mass of dissolved chromium in
groundwater of about 98 percent, according to
investigators. This reduction can be seen by
Cr(VI) iso-concentration plots with time in
Figures 3-2 and 3-3. This significant reduction
in Cr(VI) in groundwater was accomplished in
approximately 2 years after 7 years of
conventional groundwater extraction and
treatment (Thomasser and Rouse, 1999;
Thomasser, 1999). Monitoring data in the June
2000 status report indicate that all but a few
groundwater monitoring wells met the 0.05
mg/L cleanup standard for Cr(VI). Levels of
Cr(VI) in 5 of the 31 monitoring wells were
slightly above the standard.
In situ remediation with chemical reductant was
terminated at this site in October 1999 and a
closure monitoring program was implemented
under EPA oversight. Groundwater extraction,
treatment, and reinjection without reductant has
continued (Thomasser, 2000a). The plan is to
shut off select recovery wells as they reach
cleanup levels for chromium (and arsenic) and
to formally close the site. However, recent
groundwater monitoring data also indicate that
sulfate concentrations in many monitoring wells
have increased to greater than the national
secondary drinking water standard of 250 mg/L
as a result of the Cr(VI) remediation effort. In
addition, manganese concentrations have
increased to greater than the standard of 0.05
mg/L in a few locations (Lau, 2000). Project
cost data were not available during preparation
of this Guide.
Figure 3-2. Plot of Cr(VI) contaminant plume at the Valley Wood Preserving site
January 1998.
25
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Figure 3-3. Plot of Cr(VI) contaminant plume at the Valley Wood Preserving site -
November 1999.
Northern California
At two former wood treating sites in northern
California, pump-and-treat techniques were
used to control off-site Cr(VI) migration prior to
initiation of in situ efforts. However, low
permeability reduced the ability to recover
groundwater. Field tests of ferrous ion injection
further reduced the permeability due to iron
precipitation and clogging of soil pore spaces.
A sulfur-based reductant was then used in a
field-scale program. Because the ability to use
groundwater extraction and reinjection was
hampered at these sites, the reductant was
introduced by direct-push pressure injection
and hydrofracturing across the plume. This
was enhanced by a program of reductant
infiltration in the vadose zone of the source
area. Groundwater monitoring at both sites has
shown a declining trend in Cr(VI)
concentrations. Treatment is ongoing. Areas
of localized elevated levels of Cr(VI) may still
exist. Additional direct-push injection was
planned for these areas prior to site closure
(Thomasser and Rouse, 1999; Thomasser,
2000b).
South Australia
This site is the location of a CCA wood-treating
facility in South Australia. The site is underlain
by cavernous limestone and is approximately
1 km upgradient of the water supply of a city of
approximately 25,000 people. Investigators are
currently involved in the in situ fixation of
chromium in both the saturated and vadose
zone.
Groundwater remediation is performed by
pumping from a series of recovery wells within
the core of the plume, treating the water by the
addition of a sulfur-based reductant, and
injecting the water into a series of wells around
the plume. During the first year of operation,
the mass of dissolved Cr(VI) in the plume was
reduced by 55 percent, despite the continued
input of chromium to the groundwater from
seepage through the vadose zone (Thomasser
and Rouse, 1999). Of this reduction, a
calculated 85 percent occurred in situ, and only
15 percent resulted from the surface treatment.
Reductions in Cr(VI) concentrations in
groundwater were most significant in areas
26
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around the margins of the contaminated
plume. The effect of groundwater treatment in
the interior of the plume has not been as
significant; Cr(VI) concentrations in monitoring
wells have been inconsistent. However,
ongoing in situ remediation is anticipated to
show reduction of Cr(VI) concentrations from
plume margins towards the interior,
progressively until the remediation target of
0.05 mg/L is achieved (Rouse, 1999a)
For treatment of the vadose zone, a series of
infiltration pits was constructed in the area of
the former wood treating cylinder sump and the
drip pad. Excess stormwater is amended with
sulfur-based reductant and discharged to the
pits. The water percolates through the solution
features, which previously served as pathways
for the contamination. The effectiveness of the
percolation is measured by a network of
pressure/vacuum lysimeters in the vadose
zone and monitoring wells in the saturated
zone. One lysimeter located in the treatment
cylinder sump yielded a sample containing 58
mg/L before the initiation of percolation. Within
2 weeks the chromium concentration in the
lysimeter sample dropped to less than 0.01
mg/L. Samples from a well near the pit
increased from a prior concentration of 3.8
mg/L chromium to a peak value of 120 mg/L
and then dropped to less than 0.01 mg/L
chromium within 10 weeks of the initiation of
percolation (Thomasser and Rouse, 1999).
South Carolina
The use of in situ geochemical fixation is also
being evaluated at the Townsend Saw Chain
site in Pontiac, South Carolina. Investigators
have been conducting on-site treatability
testing in order to complete the remedial
design for application of the technology.
Discharged waste water from spent chromium
plating solutions has leached Cr(VI) into the
shallow groundwater aquifer. Soils in localized
areas of the former waste ponds have Cr(VI)
concentrations ranging from a few mg/kg to as
high as 30,000 mg/kg. Groundwater impacts
downgradient from the former ponds reach
levels approaching approximately 4 mg/L
Cr(VI). Treatability testing has been conducted
on soil and groundwater. The objective of soil
testing was to determine whether surface
application of ferrous sulfate solution would
effectively reduce Cr(VI) and immobilize
chromium in vadose zone soil. The surface
application of reductant solution was
accomplished using an existing spray field
piping network and soaker hoses. The average
Cr(VI) reduction percentages were 84 percent
for saturated soils, and 61 percent for drier,
unsaturated soils. Reductions as high as 97
percent were measured. Although significant
reductions were observed, only one of the ten
sample locations from the first two applications
met the soil remediation goal of 16 mg/kg
established for this site. (Pretreatment surface
soil concentrations ranged from 39 to 1,500
mg/kg, while post-treatment concentrations
ranged from 17 to 680 mg/kg. For soil samples
collected from a depth of 1.5 to 2 feet below
grade, pre- and post-treatment concentrations
of Cr(VI) were 7.9 mg/kg and 3.2 mg/kg,
respectively.) The remediation goal of 16
mg/kg is currently being re-evaluated by
Region IV EPA and may be increased (ABB,
1998a).
Two separate treatment tasks and two full-
scale injections related to in situ groundwater
remediation have been conducted at the site
thus far. The first treatment task consisted of
a small-scale injection of ferrous sulfate
reagent into a single vertical injection well in
the area where the original source of
groundwater contamination was located.
Because the first treatment task was situated
in an area having relatively low Cr(VI)
concentrations (0.2 to 0.3 mg/L), a second
treatment task, also consisting of a small-scale
injection of ferrous sulfate reagent into a single
vertical injection well, was conducted in an
area of highest groundwater contamination [2
to 4 mg/L of Cr (VI)]. The data obtained from
the first two groundwater treatment tasks were
used to design the first full-scale injection,
which was situated along the upgradient edge
27
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of the plume. Data obtained from the first full-
scale injection was then used to plan the
second full-scale injection, which was situated
approximately 200 feet downgradient of the
first injection.
In the first treatment task, an acidic ferrous
sulfate reductant solution was introduced into
the subsurface via the vertical injection well
using the existing sprayfield piping network and
chemical feed system to evaluate its effect on
groundwater at the Townsend Saw Chain site.
A chloride tracer was also utilized during the
test to monitor the progression of the reagent
plume. The injection of ferrous sulfate solution
into the aquifer plume was successful in
reducing total and Cr(VI) concentrations in
groundwater to below the remediation goal of
0.1 mg/L. After treatment, initial groundwater
concentrations up to 0.38 mg/L were effectively
reduced to less than 0.04 mg/L.
Concentrations of sulfate, chloride, and total
and ferrous iron were elevated during
treatment due to their introduction into the
aquifer but steadily returned to pretreatment
levels after treatment. The concentration of
total chromium initially increased in the aquifer
during treatment, most likely due to the
displacement of sorbed Cr(VI) ions from the
aquifer matrix by sulfate ions. This test was
limited to groundwater containing relatively low
concentrations of Cr (VI), between 2 and 3
times the treatment standard of 0.1 mg/L. A
second injection test was planned in an area
expected to contain the highest concentrations
of Cr(VI) to evaluate treatment effectiveness in
high-concentration areas (ABB, 1998b).
The second treatment task utilized one vertical
injection well and a network of six monitoring
wells. Overall, only limited data were obtained
during the second treatment task, as compared
to the first task, due to an unexpected flow
path observed in the vicinity of the extraction
well. The reagent plume did not follow the
expected linear path to the extraction well, but
rather followed an arched path around the
downgradient monitoring wells. However, the
data that was collected provided confirmation
that Cr(VI) was again successfully treated to
below the remediation goal. Results also
indicated that approximately 10 feet of vertical
dispersion was achieved.
Based on data generated during the two,
small-scale, treatment tasks and an
engineering evaluation that identified the most
effective and feasible method of injection, the
initial phase of full-scale treatment of
groundwater was designed and constructed.
Data from the small-scale tasks indicated that,
at a minimum, a 20-foot spacing for injection
wells would effectively provide for lateral
dispersion of the reagent. However, to provide
the most feasible alternative for injection, a 40-
foot spacing of vertical injection wells was
preferred. To evaluate the effectiveness of
lateral dispersion using a 40-foot spacing,
ferrous sulfate solution was initially only
injected into every other injection well during
the first full-scale injection. After the first six
weeks of post-injection monitoring, data
suggested that injection at 40-foot centers was
not effective in dispersing the reagent.
Therefore, injection into the remaining wells
was initiated to provide data to complete an
evaluation of full-scale treatment effectiveness.
However, in the same week that the second
injection was initiated, a trend of increasing
sulfate concentrations in the first four
dispersion wells was confirmed. The
increasing sulfate concentrations in dispersion
wells indicated that the reagent plume,
although exhausted (i.e., lacking ferrous iron),
had successfully reached the dispersion wells.
Thus, the 40-foot spacing of injection wells
appeared to be effective in laterally dispersing
the reagent into the aquifer. Although the
injection was successful in laterally dispersing
the reagent, effective treatment over the entire
width of the treatment cell was not
accomplished due to the exhaustion of the
reagent. The exhaustion appeared to be due
to the buffering capacity of the aquifer in the
vicinity of the injection wells. According to EPA,
it was later determined that the buffering
capacity in this location had probably been
altered due to the past application of high pH
effluent water via spray field from an
electrochemical precipitation treatment system
28
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for pumped contaminated groundwater.
Therefore, it is anticipated that the
downgradient aquifer buffering capacity will not
be as high (will be normal for the site) and
reagent exhaustion will not be a significant
problem for the rest of the site (Howard, 2000).
A second, full-scale injection was then
conducted to attempt to overcome the
buffering capacity of the aquifer. During the
second injection, a larger volume of acidified
reagent was injected into the aquifer (over a
30-day period). The pH of the injection
solution was also lowered from approximately
3.0 to 2.5 to help overcome the buffering
capacity. Unfortunately, during the second
injection, the aquifer continued to buffer the
reagent, causing the iron to precipitate.
Furthermore, due to the mass of iron
precipitate accumulating in close proximity to
the injection wells, back pressure within each
well increased significantly (Tremaine, 1999).
Iron precipitation and clogging is a potential
complication of treatment using a ferrous
sulfate-based reductant. This second, full-scale
injection was later repeated (August, 1999) at
a downgradient location. Data collected over
four weeks of monitoring indicates that
effective treatment occurred at a distance of at
least 37 feet downgradient of the injection line,
and that good lateral dispersion occurred.
Based on this data, it is expected that with
additional injection of reductant, effective
treatment at the 100-foot distance will be
achieved (Harding Lawson Associates, 1999).
The remedial design for surface soils and
groundwater treatment at this site is based on
the treatment tasks conducted during the
remedial planning phase. For surface soils,
approximately 2,400 yd3 of soil containing
Cr(VI) above the current remediation goal of 16
mg/kg were identified in the ROD for the site.
This surface soil may be treated using surface
application of reductant. However, EPA is
reviewing the current remediation goal and
may elect to change it to a recently proposed
value of 144 mg/kg. This value is based on
evidence developed during the remedial design
which does not show the expected leaching of
Cr(VI) to groundwater. EPA is preparing an
Explanation of Significant Difference that will
incorporate the revised remediation goal
(Howard, 2000). If the higher value of 144
mg/kg is approved, a much smaller volume of
surface soil will be contaminated above this
level. In this case, approximately 60 yd3 of
contaminated surface soils may be excavated
and disposed off-site (Harding Lawson
Associates, 2000).
Groundwater will be treated in situ by injecting
reductant into a series of injection wells placed
along a line that transects the width of the
plume. Injections will start at the upgradient
portion of the plume and proceed in the
downgradient direction. Injections along each
subsequent injection line will be initiated after
monitoring data from the upgradient injection
indicate that the entire treatment cell will be
effectively treated. Injection lines will be
spaced to ensure capture of any ferrous
sulfate reagent by the existing extraction well
network and minimize the potential for
contamination of a seep area and
downgradient surface water bodies. It is
anticipated that high ferrous iron (150 to 250
mg/L) and sulfate concentrations (700 mg/L),
and low pH will occur as a result of chemical
treatment. Therefore, capture of treated
groundwater and excess reagent will be
important to prevent any negative off-site
impacts. Sulfate is expected to naturally
attenuate to concentrations below the national
secondary drinking water standard of 250
mg/L. However, contingencies including
predicting potential breakthrough by modeling
mass loadings, and off-site groundwater
monitoring for early warning of sulfate
breakthrough will be utilized to help prevent
off-site contamination of surface water
(Harding Lawson Associates, 2000).
Based on an evaluation conducted during
treatment tasks, the potential for re-oxidation of
Cr(lll) at the site appears to be minimal.
However, to ensure that re-oxidation of Cr(lll)
will not occur, several studies will be
undertaken during the remediation to assess
the long-term stability of the chromium
reduction (Harding Lawson Associates, 2000).
29
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Other potential risks or uncertainties for
groundwater remediation using ferrous sulfate
reductant at this site, and their proposed
resolution, are as follows (Harding Lawson
Associates, 2000):
Small-scale heterogeneities may be present
within the treatment cells. Lithologic data will
be used to evaluate the potential presence
of any areas of gross heterogeneity (e.g.,
areas of high clay content). Flow rates and
injection pressures will be managed
according to general permeability
characteristics of the aquifer media.
Pre-acidification will be conducted and the
pH drop will be monitored prior to and
subsequent to each ferrous sulfate injection.
Areas of high buffering capacity where the
pH is difficult to adjust may require
additional pre-acidification prior to treatment.
Iron fouling may occur in or in the vicinity of
injection wells. Pre-acidification prior to
ferrous sulfate injection, and quality control
of the injected reagent, will be employed to
help reduce or prevent iron fouling. The use
of larger diameter (2-in) injection wells
should also reduce iron fouling in the wells.
"Clogging" of the aquifer from the formation
of the solid iron-chromium hydroxide
complex may decrease the yield and supply
of make-up water from on-site extraction
wells. Other source areas of groundwater
upgradient of the spray fields, or a potable
water supply, may be used for make-up
water.
Implementation of the full-scale groundwater
remedial action began in late 1999 and is
currently ongoing. Preliminary monitoring data
indicate that the treatment process is effectively
reducing Cr(VI) concentrations to less than the
remediation goal (MCL) of 0.1 mg/L in most
locations. Some monitoring locations are not
being effectively treated and have Cr(VI)
concentrations above the MCL. This
contamination appears to be spatially limited and
may require spot-treatment to fully remediate
(Harding Lawson Associates, 2000). As
expected, sulfate levels in groundwater have
increased due to the injection of ferrous sulfate
reagent. Data are not yet available to determine
if post-treatment sulfate levels return to pre-
injection levels or if sulfate migration outside the
treatment capture zone(s) is occurring.
It has been estimated that approximately
563,000 cubic yards of saturated zone material
requires remediation of groundwater to below
the regulatory limit of 0.10 mg/L Cr(VI). The
remedial action cost estimate for the
groundwater remedy at this site is $1,677,800
(Harding Lawson Associates, 1999). This
results in a cost of approximately $3 per cubic
yard, but does not include costs for the
preliminary treatment tasks. A cost estimate for
treatment of surface soils was not available
since the remedy had not been finalized.
3.2 PRBs (Treatment Walls)
3.2.1 Technology Description
PRBs provide in situ treatment of groundwater
emitting from source zones (Vance, 1997). They
are designed as preferential conduits for
contaminated groundwater flow (U.S. EPA,
1997a). These reactive barriers differ from the
highly impermeable barriers, such as grouts,
slurries, or sheet pilings, which have previously
been used to restrict the movement of
chromium-contaminated groundwater plumes.
PRBs can be installed as permanent, semi-
permanent, or replaceable units across the flow
path of a contaminant plume and act as a
treatment wall. Natural hydraulic gradients
transport contaminants through the strategically
placed reactive media (U.S. EPA, 1996b).
When the contaminated water passes through
the reactive zone of the barrier, the
contaminants are either immobilized or
chemically transformed to a more desirable
(e.g., less toxic, more readily biodegradable,
etc.) state (U.S. EPA, 1997a). In the case of
chromium, it is immobilized by precipitation onto
reactive media or aquifer solids. PRBs are not
currently used to directly remediate contaminant
source areas, only to intercept and treat
30
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contaminant plumes. The installation, design,
and reactive media for PRBs are discussed in
the following text.
PRBs are installed downgradient of a source
zone, vertically intersecting the contaminated
groundwater flow. They can be installed with
trenching, if the targeted portion of the aquifer is
shallow and surface improvements do not
interfere with access. They can also be installed
by well injection. Injection through standard
vertical wells is the least expensive option but
horizontal borings can be installed beneath
existing structures and are able to create a
uniform reactive zone. This is more difficult to
achieve through vertical wells (Nyer and Palmer,
1997). Most health and safety issues regarding
PRBs are mainly associated with installation of
the wall, and will vary according to the method of
installation used. Environmental impacts from
treatment wall installation, as well as
maintenance, may be less than with other
technologies due to the placement of all
treatment materials underground, with minimal
disturbance to surface activities (Vidic and
Pohland, 1996).
Two basic designs are being used in full-scale
implementation of reactive barriers: (1) funnel-
and-gate, and (2) continuous trench. These
designs are depicted in Figures 3-4a and 3-4b
(U.S. EPA, 1997a). Basically, for the funnel-and-
gate system, an impermeable funnel, typically
consisting of interlocking sheet pilings or slurry
walls, is emplaced to enclose and direct the flow
of contaminated water to a gate or gates
containing the permeable zone of reactive
media. Due to directing large amounts of water
through a much smaller cross-sectional area of
the aquifer, groundwater velocities within the
barrier will be higher than those resulting from
the natural gradient. Sheet pilings eliminate the
removal of soil and reduce the soil disposal cost.
Depending on the type of slurry wall, some
portion of the excavated soil may be
incorporated into the wall; however, soil disposal
costs must be taken into account for these types
of funnels. Due to aquifer heterogeneity and to
minimize groundwater mounding, a low funnel-
to-gate ratio is preferred. In order to assure
complete capture of the plume, the length of a
funnel-and-gate system is typically 1.2 to 2.5
times the plume width depending on the funnel-
to-gate ratio and the number of gates.
The funnel-and-gate configuration also allows
for reactive material to be more easily replaced;
however, field experience indicates funnel-and-
gate is more susceptible to hydraulic
uncertainties which cause bypass of flow
around the system (O'Hannesin, 1999). The
continuous trench is simply a trench that has
been excavated and simultaneously backfilled
with reactive media, allowing the water to pass
through the barrier under its natural gradient
(USEPA, 1997a). A continuous trench
distributes the reactive material across the
entire path of the contaminated groundwater.
Of the alternatives, this configuration is least
sensitive to complexities in the flow field and
does not significantly alter the natural
groundwater flow path (O'Hannesin, 1999).
Installation
In more shallow areas (less than 35 ft)
conventional excavation and replacement
methods are typically utilized for PRB
installation. These methods are typically less
expensive than those implemented deeper into
the subsurface. For deeper installation,
excavation and replacement can be costly and
are often influenced by the need to excavate to
considerable depths through uncontaminated
soil before reaching the plume. Several
construction methods are available to
accommodate the various configurations at
shallow depths. The least expensive trenching
method is backhoe trenching, which can be
implemented if the formation soil does not cave
in; however, the limitation is the excavation
width. A continuous trenching machine, which
is currently limited to depths of less than 35 ft,
allows for simultaneous excavation and
backfilling without an open trench, and it allows
for very rapid installation; however, it has
31
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of PRB
Treated Water
Figure 3-4a Plume capture by a funnel-
and-gate system. Sheet piling funnels
direct the plume through the reactive
gate.
considerably larger equipment and slightly
higher mobilization costs. Another common
installation method is cofferdam or sheet pile
excavation boxes that are formed and braced
using interlocking sheet piling. The sheet piling
maintains the dimensions of the treatment zone
during excavation and backfilling. After
backfilling is complete, the sheet piling is
removed and the groundwater is allowed to flow
through the treatment zone. Trench boxes, like
sheet piling, are used to maintain trench
integrity during excavation and backfilling
operations. Auger holes have also been utilized
to install treatment zones. Rotating a continuous
flight of hollow stem augers into the required
depth, the reactive material can be placed
through the auger stem as the augers are
removed. These treatment zones can be
created by overlapping holes or in well arrays
where two or more rows are required. Methods
for deeper installations are also being used or
researched and evaluated. Caisson installation
involves driving a large circular steel caisson
into the ground and augering out the native
material. The caisson is then backfilled with a
reactive material and removed. Overlapping or
tangential caisson emplaced treatment zones
can be used to create a larger permeable
treatment zone. However, the overlapping
caissons cause wastage of iron ranging from
Figure 3-4b Plume capture by a
continuous trench system. The plume
moves unimpeded through the reactive
gate.
10 percent to as high as 30 percent. A mandrel
or H-beam is a hollow steel beam with a
disposable shoe at the leading edge that is
driven into the ground to create a thin continuous
treatment zone. Once the mandrel reaches the
maximum depth of the treatment zone, the
reactive material is placed inside the mandrel
and the disposable shoe is removed. This
process is then repeated creating a continuous
zone of reactive material. In previous
applications, parallel treatment zones were
created to provide sufficient reactive material and
to reduce the risk that the contaminants would
not come in contact with the reactive material.
Another installation method that has been
proposed at several sites, but has not been used
to date, is ground freezing which has been
implemented in the construction industry for
many years, and involves the use of refrigeration
to convert in situ pore water into ice. The ice
acts as a bonding agent, which fuses together
particles of soil to increase the strength of the
mass, and makes it impervious. Excavation can
be performed safely inside the barrier of water-
tight frozen earth with conventional excavation
equipment. One deep installation method
requires that the reactive materials be carried in
a biodegradable slurry (bioslurry), usually guar.
32
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This method has been employed in the
construction industry for years, and is currently
being modified to implement PRBs along with
various reactive materials deep into the
subsurface. As part of pre-construction
activities, tests to determine that the site water
chemistry is compatible with the reactive
material and bioslurry mixture, and to assure
that the bioslurry breaks down over a suitable
time period at groundwater temperatures,
should be undertaken. Additional testing should
also be undertaken to determine the
effectiveness of the reactive material once it
has been in contact with the bioslurry
(O'Hannesin, 1999). A bioslurry trench
installation is similar to constructing a
conventional impermeable slurry wall. As the
trench is excavated, bioslurry provides stability
to the trench walls, and the reactive material is
placed via a tremmie tube into the trench.
Minimal contact should be made between the
reactive material and bioslurry (O'Hannesin,
1999).
Vibrated beam technology has been used for
years to install thin impermeable slurry walls
and recently has been adapted to inject reactive
material and bioslurry. The large I-beam is
driven into the subsurface and as the beam is
vibrated out, a reactive material and bioslurry is
pumped into the formation, filling the void
created by the beam. This process is then
repeated and several lines in parallel can
provide the required amount of reactive material
(O'Hannesin, 1999).
Deep soil mixing has been suggested where the
reactive material is mixed with biodegradable
slurry and pumped to the mixing augers while
they are being advanced slowly through the soil.
Over time the bioslurry breaks down allowing
the groundwater to flow through the reactive
material and aquifer mixture or treatment zone.
High costs are associated with mobilization and
demobilization for deep soil mixing (O'Hannesin,
1999).
Two other deep installation methods that have
been successfully demonstrated in the field are
jetting and vertical hydrofracturing. Jetting uses
high pressure to inject fine-grained reactive
material into the natural aquifer formation. The
jetting tool is advanced into the formation to the
desired depth, then the reactive material and
bioslurry are injected through the nozzles as the
tool is withdrawn. Either columnar zones or
diaphragm walls can be created. Vertical
hydrofracturing uses a specialized tool to orient
a vertical fracture and initiate the fracture
process. The tool is placed to the desired depth
through a borehole and the interval for fracturing
is isolated by packers. The reactive material and
bioslurry are then pumped under low pressures
into the formation to form a thin vertical plume
along the line of the induced vertical fracture
(O'Hannesin, 1999).
Site Characterization
In order to successfully install a PRB, a thorough
site characterization must be conducted. The
entire plume must flow through and react with
the reactive media. It must not be able to pass
over, under, or around the barrier and the
reactive zone must be capable of reducing the
contaminant to concentration goals without
rapidly plugging with precipitates or losing its
reactivity. To achieve this success, knowledge is
required of:
• Plume locations
• Plume direction
• Contaminant concentrations
• Hydrologic changes with time
• Concentration attenuation over time and
distance
• Stratigraphic variations in permeability
• Confining layers
• Fracturing
• Aqueous geochemistry
33
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Barrier design, location, emplacement
methodology, and estimated life expectancy are
based on the site characterization information;
therefore, faulty information could jeopardize
the entire remedial scenario.
Monitoring for regulatory compliance and
treatment performance are both necessary
when using PRBs. Compliance monitoring
determines whether regulatory contaminant
concentrations are being met.
Reactive Media
Most permeable reactive barriers use reactive
iron metal to treat chromium waste. Cr(VI) as
chromate has been shown to be reduced by
zero-valent reactive iron (Fe°). The Fe° donates
the electrons necessary to reduce the chromate
and becomes oxidized to Fe(ll) or Fe(lll). When
iron is present, the Cr(lll) can precipitate as a
mixed chromium-iron hydroxide solid solution,
which has a lower solution equilibrium activity
than pure solid-phase hydroxide. Therefore,
both the toxicity and mobility of chromium are
greatly decreased when it is reduced from
Cr(VI) to Cr(lll) (U.S. EPA, 1997a). This
reduction can be described by the overall
reaction:
Cr(VI) + 3Fe(ll) -*• Cr(lll) + 3Fe(lll)
This reaction appears to be appropriate for pH
values less than 10 and for phosphate (PO4)
concentrations less than 0.1mM. Above pH 10,
the rate of oxidation of the ferrous iron by
dissolved oxygen is greater than the rate of
oxidation of ferrous iron by CrO4 (Walker,
1999).
Work at one contaminated site has shown
sodium dithionite is capable of reducing Fe(lll)
to Fe(ll), which in turn reduces Cr(VI) to Cr(lll).
Experiments have also shown that the half-life
of dithionite is 2 to 3 days in the site's confined
aquifer. This half-life is adequate for reducing
the contaminants in the plume, while ensuring
that dithionite does not remain as a contaminant
in the groundwater for an extended time (U.S.
EPA, 1995a). This discovery has been one of
several steps
towards more effective utilization of Fe° PRBs in
chromium remediation.
Dithionite is a sulfur-containing oxyanion which
breaks down quickly in aqueous solution to form
two sulfoxyl radicals. These radicals react
rapidly to reduce ferric iron in minerals and
oxides which occur naturally in most aquifer
sediments. Amonette et al. (1994) have shown
that, within the aquifer, the injected dithionite
reacts with structural iron in oxyhydroxide and
iron-bearing layer silicate mineral phases,
reducing Fe(lll) to Fe(ll) according to the overall
reaction described by Equation 1:
(1) S2O4"(aq) + 2Fe(lll)(s) +
2SO2"(aq) + 2Fe(ll)(s) + 4H+
The reduced sediments in the treatment zone
can remove redox-sensitive contaminants from
groundwater flowing through the zone. Within
the zone of dithionite-reduced sediments,
aqueous chromate reacts with Fe(ll) produced by
the dithionite reaction (Equation 1) and is
precipitated as a solid hydroxide (e.g. Cr(OH) )
according to the example reaction described in
Equation 2:
(2) HCrO"(aq)+ 3Fe(lll)(s) + 4H+^D
Cr(OH)3(s)
The majority of PRB remediation techniques
discussed at symposiums today focus on the use
of granular iron filings (Fe°) as the reductant of
choice (Cercona, 1995). Fe° can also be used to
dechlorinate trichloroethene (TCE) (U.S. EPA,
1997a).
Iron filings are relatively inexpensive and are
available in coarse particulate sizes that result in
packing densities less than 50 percent of
theoretical; therefore, it is easily shipped to sites.
The use of these materials in particulate form
presents materials-handling difficulties,
particularly when the process is being conducted
in situ and an absorbent material must be
removed from the ground after it has become
fully "loaded" with the contaminant.
34
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A ceramic foam process has an advantage over
packed iron filings due to its monolithic nature.
The blocks of foam can be easily handled and
mounted in a frame or cassette or strung like
beads on a wire to facilitate placement and
removal. Densities of the foamed material can
be adjusted to increase compressive strength to
withstand geostatic loads (Cercona, 1995).
Foams are made with relatively high surface
(Bowman et al., 1999). Natural zeolites can be
treated with cationic surfactants to alter their
surface chemistry and improve their affinity for
sorption of anionic metals such as chromium. In
addition, zeolites can be combined with Fe° to
create the combined effect of adsorption and
reduction. This may lead to a decrease in the
amount of material required to achieve a given
level of Cr(VI) removal (Zhaohui, 1999).
Table 3-3. Comparison of Granular vs. Foam Iron Reactive Media for PRBs
Size/Packing Density
Surface Area
Raw Material/Unit Cost
Typical Excavation Cost
Granular Iron
20-40 mesh
42 to 45 percent
212lbs/ft3
1 .1 m2/g specific surface area
1 05,870 m2/ft3
$75/ft3
1 ,400 m2/$
$500,000 for 2,000 ft3 gate
10 to 15 percent of theoretical
60 to 75 Ibs/ft3
Greater than 4 m2/g
Greater than 120,000 m2/ft3
$1/lb
$60 to $75
1 ,600 to 2,000 m2/$
Low
concentrations with metallic iron contents
between 92 and 94 percent and high
specific surface areas exceeding 5 m2/g.
The development of foams has focused on
maximizing surface area, iron purity content,
and the possible incorporation of secondary
materials that could improve the rate of
adsorption/immobilization of heavy metals.
Table 3-3 shows a comparison of granular
iron and iron foam (Cercona, 1995). The
use of granular iron has been more typical
for PRB applications.
Zeolites are also being evaluated as an
adsorbent material for use in PRBs. Zeolites
have large specific surface areas, high
adsorption capacities, high CECs, good
hydraulic characteristics, and relatively low
cost
Biotic applications such as biomineralization
can also be accommodated with the PRB
approach to treatment of chromium
contaminated groundwater. A
biomineralization application is discussed in
Section 3.3.
Advantages
• Actual in situ contaminant remediation.
• Passive remediation, no ongoing energy
input and limited maintenance following
installation; reduced operation and
maintenance (O&M) costs compared to
pump-and-treat.
• No required surface structures other than
monitoring wells following installation.
• Can remediate plume even when the source
of the plume cannot be located.
35
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• Should not alter the overall groundwater
flow pattern as much as high-volume
pumping.
• Contaminants are not brought to the
surface; i.e., no potential cross-media
contamination.
• No disposal requirements or disposal costs
for treated wastes.
• Avoids mixing of contaminated and
uncontaminated waters that occurs with
pumping.
• Foams have the capability to be a
"customized" system.
• Foams have a very high specific surface
area.
• Foams have a tailored pore size and high
permeability/porosity, which controls mass
transport capabilities.
• Foams have a controlled composition.
• Foams have the potential for easy retrieval
(if necessary) and if they are retrieved, their
volumes can easily be reduced for
disposal.
Limitations
• Currently restricted to shallow plumes,
approximately 50 ft or less below ground
surface.
• Plume must be very well characterized and
delineated, for example, no fractured rock
or excessive depth to contaminant plume.
• Limited long-term field testing data is
available and field monitoring is in its
infancy.
• Limited field data concerning longevity of
wall reactivity or loss of permeability due to
precipitation.
• Currently no field-tested applications to
remediation of contaminant source areas.
• Blockage of the pore space with products of
reaction processes, particularly with
injection based systems.
• Does not allow the degree of aquifer
hydraulic control of active approaches like
pump-and-treat.
3.2.2 Status
EPA recognizes PRBs as a technology with the
potential to more effectively remediate
subsurface contamination at many types of sites
at significant cost savings compared to other
more traditional approaches. It is actively
involved in the evaluation and monitoring of this
new technology to answer questions regarding
long-term system performance, and to provide
guidance to various stakeholder groups (USEPA,
1997a).
The Waterloo Centre for Groundwater Research
has developed in situ PRBs for treatment of
inorganic contaminants, including chromium, at
several sites in Canada and the U.S. Inorganic
contaminants in groundwater are treated using in
situ porous reactive walls.
Treatment wall materials are placed in the path
of the plume and react with the contaminant via
reduction and precipitation. Preliminary results of
field trials indicate contaminant concentrations
were decreased by orders of magnitude and to
below drinking water levels (U.S. EPA, 1995a).
As of April 1997, there were 124 treatment wall
projects identified for all classes of site
contaminants. Fe° was employed as the reactive
media in 45 percent of the 124 projects. For
inorganic contaminant projects, 31 percent
involved the remediation of chromium (Sacre,
1997).
36
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Figures 3-5 and 3-6 show the breakdown of and the types of projects addressing chromium
inorganic contaminants addressed by PRBs remediation (Sacre, 1997).
Arsenic 9%
Cadmium 9°=
31%
7%
'. 3%
Copper 3%
* /: 3%
Selenium 4%
Other 9%
Figure 3-5 Breakdown of inorganic contaminants addressed by PRBs.
14-
12-
10-
8 -
6 —
4-
2-
x-
x
x
X
X
X
X
9
I
I
P"
0
3
x_
5
.xf—
_^ |
„„„„„„„„„„„„„
r7!
j__
I I _ : I
Lab Pilot Field Commercial
Figure 3-6 Breakdown of types of PRB projects addressing chromium remediation.
37
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A Permeable Barriers Action Team (PBAT)
was established in March 1995 as part of
EPA's Remediation Technologies
Development Forum (RTDF). The mission of
the PBAT is to accelerate the development of
cost-effective PRB technologies. Its efforts
focus on defining the hydraulics, geochemistry,
and reactions occurring in the media and
aquifers; demonstrations that validate the
technology's effectiveness; protocols for
design and implementation; effective
emplacement technologies and configurations;
economic analysis of treatment costs; and
public and regulatory acceptance of
technology (USEPA, 1996b).
3.2.3 Performance and Cost Data
The costs of the impermeable sections of the
treatment wall system can be obtained from
experiences with slurry walls or sheet pile
installations. If the reactive media is Fe°, the
cost of the media can be estimated based on
the density of about 2.83 kg/m3 and a cost of
approximately $350 to $450/ton. A 1996 review
suggested that installation costs between
$2,500 and $8,000 per L/min of treatment
capacity can be used as a rule-of-thumb for
estimating the capital cost of these systems.
Since zero-valent treatment walls is a patented
technology, a site licensing fee, which has
been typically 15 percent of the capital costs
(materials and construction costs), may also be
required (Vidic and Pohland, 1996).
Using continuous trenching machine
installation, costs range from $200 to $400 per
linear foot (Puls, 2000). The sheet pile
excavation boxes average about $80 ft2 and
trench boxes range about $10 to $20 per ft2.
Caisson implementation averages about $200
per vertical foot, and a mandrel installation
ranges from $10 to $20 per ft2. A bioslurry
trench can be installed for $15 to $25 per ft2.
The vibrated beam method can be installed on
the order of $10 per ft2 and deep soil mixing
cost can range from $75 to $120 per yd3. To
utilize the jetting method may cost on the order
of $75 per vertical foot (O'Hannesin, 1999).
A principal advantage of PRBs over other
groundwater remediation approaches is
reduced O&M costs. Other than groundwater
monitoring, the major factor affecting O&M
costs is the need for periodic removal of
precipitates from the reactive media or periodic
replacement or rejuvenation of the affected
sections of the permeable wall. O&M costs
between $1.3 and $5.2 per 1.000L of treated
water can be used as a rule-of-thumb for
estimating the O&M costs of these systems
(Vidic and Pohland, 1996).
A cost analysis conducted by Manz and Quinn
(1997) indicates that use of PRBs can result in
significant cost savings over a comparable
groundwater extraction and treatment system.
In their study of two sites, they indicate that
while capital costs vary, the annual estimated
operation and maintenance costs for a
treatment wall were between $20,000 and
$27,120, as compared to between $55,000 and
$100,000 for a pump-and-treat system.
However, actual cost savings depend on the
initial capital costs for the barrier installation and
the estimated longevity of the reactive barrier.
Depending on the scale of analytical monitoring
required, operational costs for a PRB may be
70 to 90 percent less than the cost of a pump-
and-treat system per year. This is due to the
fact that no provision is necessary for the
disposal of recovered water and the system is
mechanically passive (Vance, 1997).
There are several sites which currently employ
PRB technology for chromium remediation.
Some of these applications are discussed in the
following text.
38
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Elizabeth City, North Carolina PRB
An integrated technical demonstration program
for chromium remediation was established at a
U.S. Coast Guard (USCG) facility in Elizabeth
City, North Carolina which operated a chrome
plating shop for 30 years. Activities at this
facility resulted in the release of chromic acid
into the soils below the shop. A detailed
characterization of the underlying soils and
groundwater of the chrome plating shop was
performed to provide information on the extent
of contamination at the site and the potential
for off-site migration and environmental impact
(Khan, 1999).
Installation of the PRB at this site involved two
phases; a pilot-scale field test and a full-scale
field demonstration (U.S. EPA, 1997). In the
pilot-scale field test, two types of reactive iron
were mixed in equal volumes with coarse sand
and native aquifer material on-site and poured
through hollow stem augers drilled to the
appropriate depth. Frequent geochemical
monitoring of groundwater in the test zone was
performed to assess the removal of dissolved
chromate from the groundwater and to confirm
or elucidate the proposed chemical
mechanisms responsible for remediation.
Results from the September 1994 field pilot
test in Elizabeth City indicated that complete
treatment of chromium in the groundwater
might be possible at this site. Chromium
concentrations at this site were reduced to less
than 0.01 mg/L, below the drinking water limit,
according to the researchers (Puls et al.,
1999).
The full-scale field demonstration consisted of
laboratory and batch column tests to determine
the granular iron mixture best suited for
simultaneously treating Cr(VI) and TCE
contaminated groundwater; three-dimensional
groundwater flow simulations to assess the
relative efficiency of a funnel-and-gate versus
a continuous wall design; design and
installation of the selected reactive barrier;
performance monitoring of the installed barrier;
and multicomponent reactive transport modeling
(U.S. EPA, 1999a, 1999b, and 1999c).
Groundwater plumes at the Elizabeth City site
contained Cr(VI) and TCE in excess of MCLs
(greater than 5 and 10 mg/L, respectively). The
release of Cr(VI) to the subsurface resulted in
the development of a well-defined plume of
groundwater containing Cr(VI) concentrations in
excess of 28 mg/L near the source. An
extensive plume of TCE overlaps the Cr(VI)
plume. Based on the preliminary studies, a full-
scale continuous wall PRB demonstration
system was installed at the same Elizabeth City,
North Carolina site on June 22, 1996. The
continuous wall design was selected because
for the Elizabeth City site, there were no
hydraulic advantages of a funnel-and-gate
design in terms of both increased capture area
and increased residence time.
The full-scale barrier was comprised entirely of
Fe°, in the form of iron filings. The reactive
media was selected based on suitable reaction
rates, desirable hydraulic properties, and lower
cost. The installed reactive barrier was 46.0 m
long, 5.5 m in depth, and 0.6 m wide. The
dimensions were selected to ensure capture of
the full horizontal extent of the Cr(VI) and TCE
contaminated plumes, and to prevent
penetration of a fine-grained geologic unit
present at approximately 8 m depth. The
reactive barrier was installed in less than
8 hours using a continuous trenching technique.
The total cost of the reactive barrier at Elizabeth
City including site assessment, design,
construction, materials, and preliminary and
follow-up work, was approximately $985,000.
The installation and granular iron costs were
estimated to be approximately $350,000. These
capital costs were estimated to be comparable
39
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to a pump-and-treat system. However, annual
O&M costs are estimated to be $32,000 for the
reactive barrier compared to $200,000 for a
pump-and-treat system (USEPA 1999a).
A detailed monitoring network was installed in
the vicinity of the PRB in November 1996.
Three rows of bundle- type piezometers with up
to 11 sampling ports, ranging in depth from
near the water table to 7 m, and 9 conventional
2-in polyvinyl chloride (PVC) monitoring wells
were utilized for performance and compliance
monitoring, respectively. Water samples were
collected from the sampling points and
immediately analyzed to determine pH, Eh,
electrical conductivity, turbidity, alkalinity,
dissolved oxygen, Cr(VI), total sulfide, and
Fe(ll). Dissolved and total inorganic
constituents were analyzed. Groundwater
monitoring was conducted seven times
between November 1996 and December 1998
(USEPA, 1999b).
The reactive barrier was designed to meet the
cleanup goal concentration of 0.05 mg/L
Cr(VI). Results from the monitoring network
established in November 1996 indicate that all
Cr(VI) has been removed from the
groundwater within the first 6 in of the reactive
barrier. Chlorinated hydrocarbons decreased
by more than 95 percent, with most multilevel
samplers showing reductions below MCLs
(USEPA, 1999). No chromium or Cr(VI) has
been detected above MCLs downgradient of
the wall either in the multilevel sampling ports
or in the compliance wells located immediately
behind the barrier. TCE concentrations were
reduced by orders of magnitude within the
barrier, although TCE concentrations of up to
15 jug/L were observed downgradient in two
compliance wells, above the MCL of 0.5 j
These elevated downgradient TCE
concentrations were possibly due to a portion
of the TCE plume going underneath the barrier
wall and therefore receiving inadequate
treatment. The pH increased from background
values of pH 6 to 7 to values of 9.0 to 10.5
within the reactive barrier, and then fell to
background values within 2 m downgradient of
the iron wall. The Eh showed a corresponding
sharp decline, from background values of 100
to 500 mV to very low values of -400 to -600
mV within the reactive barrier. The Eh rose
again downgradient of the reactive barrier. Low
Eh and high pH values within the reactive
barrier indicate that conditions were suitable for
the reduction of Cr(VI), the precipitation of Cr(lll)
oxyhydroxides, and the reductive dechlorination
of TCE. The alkalinity values decreased from
background values between 40 and 100 mg/L
(as calcium carbonate - CaCO3) to <10 mg/L (as
CaCO3) within the reactive barrier, which may
be the result of carbonate mineral precipitation
(U.S. EPA, 1999b).
The precipitation of secondary minerals within
the barrier may have an important impact upon
the long-term performance of the barrier. Slug
tests performed in February 1997 indicate that
the hydraulic conductivity of the granular iron
was still significantly greater than the hydraulic
conductivity of the aquifer. Reactive transport
modeling was used to look at the precipitation
of secondary minerals and their effect on long-
term performance efficiency, as well as other
aspects of treatment. The model results indicate
that over a long period of time, porosity may
decrease significantly, which will almost
certainly affect the hydraulic properties of the
treatment system. The reactivity of the
treatment material may also decline over time
which could reduce the contact time of the
contaminants with the treatment material. This
may lead to the imcomplete treatment of
contaminants that require a long residence time
(U.S. EPA, 1999c).
40
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Four iron foam samples with various
compositions, densities, and specific surface
areas were also tested at the Elizabeth City
site. Two of the samples showed positive
results for chromium removal. Following these
results, additional samples were requested for
more specific evaluations. It was found that the
inclusion of aluminosilicate materials helped
control the reaction kinetics by buffering the pH
of the groundwater associated with the iron
reduction reaction. If the pH of the system was
allowed to approach 9.5 to 10.0, the reaction
kinetics were slowed. When aluminosilicates
were present, the pH of the system remained
under 8. The additional samples being tested
at this site were used to determine the benefits
of an iron foam containing the aluminosilicate
(Cercona, 1995).
Hanford, Washington In Situ Redox
Manipulations (ISRM)
Following years of weapons production at the
U.S. Department of Energy (DOE) Hanford site
in south central Washington State, numerous
groundwater plumes are currently impacting
the Columbia River. Pump-and-treat systems
are operating at two of the plumes. Because
PRBs have proven elsewhere to be more
effective than these methods, DOE opted to
test these on a third plume. Since the plumes
average about 80 ft below the surface, ISRM
was used as an alternative to "trench and fill"
because it can be applied through
conventional groundwater wells. The ISRM
approach extends the permeable treatment
zone concept to sites where the groundwater
contaminant plumes are too deep to be treated
by excavation or by trench-emplaced
permeable barriers (Fruchter, 1999).
ISRM using sodium dithionite (or other
reagents) creates a reducing environment for
reduction of Cr(VI) to Cr(lll). The goal of the
ISRM method is to create a permeable
treatment zone in the subsurface to
remediate redox-sensitive contaminants. Redox
sensitive contaminants in the plume are
immobilized or destroyed as they migrate
through the manipulated zone. A permeable
treatment zone is created by reducing the ferric
iron in the aquifer sediments to ferrous iron.
The treatment zone is created by injecting
appropriate reagents and buffers (e.g., sodium
diothonite and potassium carbonate) to
chemically reduce the structural iron in the
sediments.
Once sodium dithionite was selected as a
preferred reagent, a variety of batch and
column experiments with sediment and
dithionite were performed by Amonette et al.
(1994). These bench-scale studies were used
to develop an understanding of the important
reactions, final reaction products (i.e.,
residuals), and nature and fate of any ions
released from the sediments and sediment
surface coatings under reducing conditions
(e.g., mobilization of trace metals).
During the summer of 1995, investigators ran a
demonstration using ISRM. Pilot-scale
experiments tested the feasibility of using ISRM
with chemical reagents using a forced gradient,
single well, reactive tracer test. Field
experiments involved injecting sodium dithionite
into an aquifer creating a 60- to 100-ft diameter
geochemical PRB ahead of a chromium plume.
Figure 3-7 depicts the ISRM treatment approach
(Cummings and Booth, 1997).
This treatment approach could also be
classified as a reactive zone technology (see
Section 3.3); however, researchers consider it
a PRB. After allowing 5 to 30 days for the
reaction to occur, water containing the reaction
byproducts and any remaining reagent was
pumped out. The experiment was designed to
evaluate the longevity of the system to maintain
a reducing environment (U.S. EPA, 1995b). A
buffered sodium dithionite solution (21,000
gallons) was injected into the test site aquifer
41
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over a period of 17.1 hours. A subsequent
residence time of 18.5 hours was allowed
before sampling and analysis of the
groundwater was conducted. No significant
plugging of the well screen or aquifer formation
was detected during any phase of the test.
Sixty to 100 percent of available reactable iron
in the aquifer sediments was reduced by the
injected dithionite.
Groundwater monitoring 10 months after the
injection showed a reducing environment
persisting within the redox area. Cr(VI) was
reduced and levels remained below the
detection limits, and total chromium levels were
below 8 mg/L and continued to decline
(Cummings and Booth, 1997). Measurements
taken after 3 years show that total and Cr(VI)
levels were both below the detection limit (<7
ppb), down from 70 ppb initially.
Cnromate Contamination
Figure 3-7 ISRM treatment system diagram.
42
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During fiscal year 1998, a five-well, field-scale
treatability test was installed at the Hanford
100D Area. The test created a PRB 150 feet
long and 50 feet in width. The barrier was
placed about 500 feet from the Columbia River
in a chromate plume with concentrations in the
1,000 to 2,000 ppb range (as chromium). The
depth to groundwater was about 85 ft. An
average of about 27,000 gallons of buffered
sodium dithionite solution was injected in each
of the five wells. The sodium dithonite
concentration averaged approximately 0.08/W.
The average reaction phase was 35 hours.
Recent monitoring data have shown that
chromate concentrations in the reduced zone
have decreased to below detection limits. In
addition, chromate concentrations have begun
decreasing in several downgradient monitoring
wells. Monitoring is continuing. In addition to
monitoring wells near the site, a series of
sampling tubes have been placed along the
bank of the Columbia River. Sediment cores
will be taken in the reduced zone during the
summer of 1999 to determine the amount of
ferric iron reduced. The results to date are
summarized in Williams et al. (1998).
Based on the success of the treatability test,
DOE decided to deploy a full-scale barrier at
the 100D Area site. Current plans call for the
expanded barrier to be approximately 1,000 ft
in length. It will be constructed at the same
site as the treatability test barrier (Fruchter,
1999).
The ISRM researchers developed a cost
comparison between ISRM and traditional
pump-and-treat methods. The cost estimates
were based on costs from the ISRM field-scale
treatability test and from actual
implementations of other pump-and-treat
systems and used similar operation and
remedial objectives. Under the conditions
established at the Hanford site and based on
a 10-year project lifetime, ISRM realized an
overall estimated cost savings of 62 percent, or
$4.6 million, over traditional pump-and-treat
(Cummings and Booth, 1997). The estimated
total cost for pump-and-treat was $8.85 million,
and the estimated total cost for ISRM was
$2.95 million. These costs estimates were
based on assumptions for costs for equipment
procurement, O&M, waste management,
system monitoring, and data analysis.
The cost savings accrued by ISRM over the
10-year duration can be attributed mainly to
the negligible operating and waste
management costs for ISRM. Operating costs
for the pump-and-treat system are continuous
starting in the second year when the plant is
running at full capacity. Waste management
costs for ISRM are based on the one-time
need to treat and dispose of approximately 1
million gallons of groundwater from the
withdrawal of dithionite reagent and reaction
byproducts from within the barrier zone. Waste
management costs for pump-and-treat are
higher because of the greater volume of
groundwater that is extracted and treated, and
because waste management costs are
ongoing for the 10-year lifetime of the project.
System design costs are expected to be higher
for ISRM than traditional pump-and-treat
because of the necessity for thorough
characterization of aquifer sediments in
addition to groundwater. It should be noted that
emplacement of the ISRM permeable barrier
and required residence time in the aquifer to
initiate conditions for the redox reactions may
take several days; the life of the treatment
barrier is expected to last for a period of years.
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Zeolite-Based PRBs
Two pilot-scale PRB tests were performed with
surfactant-modified zeolite (SMZ) material by
researchers at the Large Experimental Aquifer
Facility of the Oregon Graduate Institute.
Natural zeolite material of different mesh sizes
was modified with a cationic surfactant to
improve its sorptive capacity and hydraulic
characteristics. The SMZ was bulk-
manufactured at a cost of about $460 per
metric ton (equivalent to $460/m3). The cost of
natural zeolite alone is $45 to $60 per ton. The
SMZ material was placed in a barrier frame in
the center of a tank and surrounded by sand to
form a simulated aquifer. Chromate-
contaminated water was then directed through
the reactive barrier (Bowman et al., 1999).
The first test showed that much of the
contaminant plume was being deflected under
and around the SMZ barrier. This was thought
to be due to low SMZ conductivity and a
partially plugged barrier frame. It was
concluded that care must be taken to eliminate
hydraulic restrictions at barrier/aquifer
interfaces in order to prevent plume deflection.
The SMZ was then replaced with a different
mesh size material and the barrier was
modified for the second test. No plume
deflection occurred in the second test. After 56
days of operation, no downgradient chromate
contamination was detected, and low
concentrations were detected in the barrier.
The estimated pilot test retardation factor for
chromate was very close to that predicted from
laboratory isotherm experiments. Therefore, it
was concluded that laboratory results can be
used to predict contaminant retardation using
SMZ for a larger scale PRB (Bowman et al.,
1999).
Laboratory tests were performed to evaluate
the overall efficiency of a combination of SMZ
and zero-valent iron (ZVI) for chromate
sorption and destruction. Zeolite/ZVI pellets
were first produced and then modified with a
cationic surfactant to increase contaminant
sorption, and, thus, the contaminant
concentration on the solid surface. Chromate
sorption/reduction tests with the SMZ/ZVI
reactive material were conducted in centrifuge
tubes. The mechanical stability of pellets under
saturated conditions was also evaluated.
Results indicate that the chromate sorption
capacity of pelletized SMZ/ZVI was at least
one order of magnitude higher than that of
zeolite/ZVI pellets. Also, compared to SMZ
pellets alone, the chromate removal capacity of
SMZ/ZVI in a 24-hour period was about 80
percent higher, due to the combined effects of
sorption by SMZ and reduction by ZVI.
Therefore, SMZ/ZVI pellets have the potential
to lower the amount of reactive material
required in a PRB to achieve a target level of
contaminant reduction (Zhaohui, 1999).
3.3 Reactive Zones
3.3.1 Technology Description
In situ reactive zones are based on the
creation of a subsurface zone where migrating
contaminants are intercepted and permanently
immobilized or degraded into harmless end
products. Reactive zones allow groundwater to
continue to flow naturally; the groundwater is
not funneled or directed into or through
subsurface barriers. Groundwater is not
extracted; it is a passive treatment system.
Reactive zones can be installed slightly
downgradient of the source area to prevent
mass flux of contaminants from migrating from
the source (Nyer and Suthersan, 1996). These
treatment zones are usually established in situ
by injecting reagents and solutions in
predetermined locations within the
contaminated groundwater plume, and allowing
them to "react" with the contaminants. A
physical subsurface "barrier" is not used as
with PRB technology. Typically, reactive zones
44
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do not directly treat the contaminant source
zone, but intercept and treat the contaminant
plume; however, source zone treatment may
be applicable in some cases such as low
permeability formations. Compared to PRBs,
reactive zones provide several advantages: no
excavation of contaminated soil is needed; the
installation and operation are relatively
cheaper; human exposure to hazardous
materials is minimized; and remediation of
deeper contaminated sites can be
accomplished (Yin and Allen, 1999).
Successful application of reactive zones
requires the ability to engineer two types of in
situ reactions: (1) between the injected
reagents or solutions and the subsurface
environment in order to manipulate the
biogeochemistry and optimize the required
reactions; and (2) between the injected
reagents, substrates, or microbes and the
migrating contaminants in order to effect
remediation. These reactions will differ
between sites and even within a site; therefore,
the major challenge is to design a reactive
zone to systematically control these reactions
under naturally variable conditions found in the
field (Nyer and Suthersan, 1996). Creation of
a spatially fixed reactive zone in an aquifer
requires proper mixing of the injected reagents
uniformly within the reactive zone.
Furthermore, such reagents must cause few
side reactions and be relatively nontoxic in
both their original and treated forms.
Creation of spatially fixed reactive zones to
achieve these reactions is very cost-effective in
comparison to treating the entire plume as a
reactive zone (Suthersan, 1997).
The mechanisms that can be used to reduce
the toxicity of heavy metals dissolved in
groundwater are transformation and
immobilization. These mechanisms can be
induced by both abiotic and biotic pathways.
Abiotic pathways include oxidation, reduction,
sorption, and precipitation. Biotically mediated
processes include reduction, oxidation,
precipitation, biosorption, bioaccumulation,
organo-metal complexation, and
phytoremediation (Suthersan, 1997).
Phytoremediation is discussed separately in
Section 3.7 of this Guide.
Design and Application
In situ reactive zones can be designed as a
curtain of injection points or multiple curtains to
intercept the moving contaminant plume at
various locations (see Figure 3-8).
A curtain can be installed slightly downgradient
of, or within, the source area to prevent the
mass flux of contaminants migrating from the
source. This will shrink the size of the
contaminant plume faster. If the duration of
remediation is a critical factor, another curtain
can be installed between the above two
curtains for further interception at the middle of
the plume (Suthersan, 1997). Contaminated
groundwater flows horizontally through the
established reactive zones, following the
natural hydraulic gradient.
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^—•" -^ .,« ^ Source Area
\
Carlain of Injection Wells at
Down gradient Edge of the Plume
in Situ Ffeaefe© Zone-
Source Area
Souiee: Siiherssn, 1S97
Curtain of Injection Wells
Down gradient of the Source Area
tn Situ Reactive Zone
Curtain erf Injsrtion Wells at
Down gradient Edge of the Plume
n Situ Reaetiw Zone
Figure 3-8 In Situ reactive zones curtain design concept (plan view).
Another approach to designing an in situ
reactive zone is to create the reactive zone
across the entire plume. The injection points
can be designed on a grid pattern to achieve
the reactions across the entire plume.
However, it should be noted that the cost of
installation of injection wells constitutes the
biggest fraction of the system cost, considering
both capital and operational costs. It is clear
that the reduction of the total number of
injection wells will significantly reduce the
system costs. Therefore, the curtain concept is
the preferred and most cost-effective approach
to implement in situ reactive zones (Suthersan,
1997).
An engineered in situ reactive zone has to take
into consideration how the target reactions will
impact the redox conditions within and
downgradient of the reactive zone, in addition
to degrading the contaminants with the
available residence time. In addition, careful
evaluation should be performed regarding the
selectivity of the injected reagents toward the
target contaminants and the potential to react
with other compounds or aquifer materials.
Careful monitoring, short-term and long-term,
should be performed to determine whether the
natural equilibrium conditions can be restored
at the end of the remediation process. In some
cases, modified biogeochemical equilibrium
conditions may have to be maintained over a
long period of time to prevent the reoccurrence
of contaminants (Suthersan, 1997).
The three major design requirements for
implementing an in situ reactive zone are: (1)
creation and maintenance of optimum redox
environment and other biogeochemical
parameters such as pH, presence or absence
of dissolved oxygen, and temperature, etc.; (2)
selection of the target process reactions and
the appropriate reagents to be injected to
achieve these reactions; and (3) delivery and
distribution of the required reagents in a
homogeneous manner across the entire
reactive zone, both in the lateral and vertical
directions (Suthersan, 1997).
46
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Site Characterization
The composition of interstitial water is the most
sensitive indicator of the types and the extent
of reactions that will take place between
contaminants and the injected reagents in the
aqueous phase. Determination of the baseline
conditions of the appropriate biogeochemical
parameters is a key element for the design of
an in situ reactive zone. This evaluation will
give a clear indication of the existing conditions
and the necessary steps to be taken to
optimize the environment to achieve target
reactions. Section 2.3 presents a number of
site characterization analytical parameters
required for chemical fixation processes. Other
biogeochemical parameters that may be
needed include:
• Dissolved oxygen
• Temperature
• Total dissolved and suspended solids
• Anions (NO3-, NCV, SO/', S")
• Fe (total and dissolved)
• Alkalinity
• Concentration of dissolved gases (CO2, N2,
CH4, etc.)
• Microbial population enumeration (total
plate count and specific degraders count)
• Any other organic or inorganic parameters
that have the potential to interfere with the
target reactions.
It should be noted that the number of
parameters that need to be included in the list
of baseline measurements will be site-specific
and will be heavily influenced by the target
reactions to be implemented within the reactive
zone (Suthersan, 1997).
Design of a reagent injection system entails an
extensive evaluation and understanding of the
hydrogeologic conditions at the site and
specifically within the plume and the location of
the reactive zones. Table 3-4 lists specific
geologic/hydrogeologic parameters required for
the design of an in situ reactive zone. Delivery,
distribution, and proper mixing of the injected
reagents are key elements to the success of
remediation within an in situ reactive zone.
Location and spacing of the injection wells and
the placement of screens within each well
(cluster) are critical to achieve complete
remediation.
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Table 3-4. Impacts of Various Geologic/Hydrogeologic
Parameters on the Design of an In Situ Reactive Zone.
Geologic/Hydrogeologic Parameter
Depth to water table
Width of contaminant plume
Depth of contaminant plume
Groundwater velocity
Hydraulic conductivity (horizontal and vertical)
Geologic variations, layering of various soil
sediments
Soil porosity and grain size distribution
Injection well depth and screen locations.
Number of injection wells.
Number of injection points within a well cluster.
Pressure injection vs. gravity feed.
Injection flow rate, residences time for the target
reactions.
Dilution of end products.
Mixing zones of reagents, extent of reactive zone.
Number of injection points within a well cluster.
Location of well screens within injection points.
Removal of end products resulting from
immobilization reactions (such as heavy metals
precipitation).
Injection of Reagents and Solutions
Injection of reagents and other solutions can be
implemented in two ways: (1) gravity feed, and
(2) pressure injection deeper into the well.
Figures 3-9 and 3-10 depict these two
approaches. Gravity feed is possible only when
the depth of contamination is very shallow.
Under gravity feed conditions, injected
reagents will tend to spread over the water
table as a sheet flow, and the mixing within the
reactive zone will be dominated by diffusion,
rather than advective flow.
When the depth of contamination is deeper,
multiple injection points may be required within
a well cluster at each injection point. The
reagent solution will have to be injected under
pressure into the injection well. Under this
configuration, mixing within the reactive zone
will be influenced by both advective and
diffusional transport of the reagents. The
concentration of the injected feed solution
should be dilute enough to avoid any
downward migration due to density differences
between the reagent and groundwater
(Suthersan, 1997).
48
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Reagent
Solution
Reagent
— . Sheet Flow
Diffosion
Shallow
Zone
Soy roe: Sulhersan, Iff?
Figure 3-9. Gravity feed of reagents when the contamination is shallow.
Reagent
Solution
Mix in g Zone
of Oeaoents
Contaminated
Zone
1997
Figure 3-10. Multiple cluster injection points when contamination is deep.
During gravity feed of the reagents, the lateral
spread of the injected solution will be
significant due to the sheet flow effect,
However, under pressure injection conditions,
downgradient migration of the injected
reagents and, thus the mixing zone could be
very narrow, depending on the hydrogeologic
conditions within the reactive zone. One way to
49
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overcome this problem is to install closely
spaced injection points. This option, even
though it is easier to implement, will
significantly increase the cost of the system.
Cyclic extraction and injection of adjoining
wells, treated as a pair, will create a wider
mixing zone downgradient of the injection wells
and will eliminate the need to install closely
spaced injection points. Extracted groundwater
can be used as the dilution water to maintain
the feed injection solution concentration
(Suthersan, 1997).
Reagent-Based Reactive Zones
Dissolved Cr(VI) can be reduced and
precipitated as chromic hydroxide [Cr(OH)3] by
the injection of ferrous sulfate solution into the
reactive zone at appropriate concentrations.
Reduction and precipitation in reactive zones
can also be accomplished using other chemical
reductants (refer to Section 3.1 on
Geochemical Fixation). Cr(VI) exists as
chromate under neutral or alkaline conditions
and dichromate under acidic conditions. Both
species react with ferrous ion to become Cr(lll)
and Fe(lll). Both Cr(lll) and Fe(lll) ions are
highly insoluble under natural conditions of
groundwater. The addition of ferrous sulfate
into the reactive zone may create acidic
conditions, and hence the zone downgradient
of the ferrous sulfate injection zone may have
to be injected with soda ash or caustic soda to
bring the pH back to neutral conditions
(Suthersan, 1997). Fe°, commonly used with
PRBs, can also be injected into the path of a
contaminated plume to effect chromate
reduction and precipitation.
Molasses-Based Reactive Zones
Injection of a carbohydrate solution such as
diluted molasses can promote the in situ
microbial reduction of Cr(VI) to Cr(lll)
(Suthersan, 1997). The carbohydrates, which
consist mostly of sucrose, are readily degraded
by the heterotrophic microorganisms present in
the aquifer, thus depleting all the available
dissolved oxygen present and causing
reducing conditions to develop. The primary
end product of the Cr(VI) to Cr(lll) reduction
process is Cr(OH)3, a form of Cr(lll), which
readily precipitates out of solution under
alkaline to moderately acidic conditions.
Cr(OH)3 precipitate is essentially an insoluble,
stable precipitate, immobilized in the soil matrix
of the aquifer (Nyer and Suthersan, 1996).
Other Biotic Reactive Zones
Unicellular Yeast: Chromium can be removed
from groundwater by the unicellular yeast,
Saccaromyces cerevisiae. Several species of
bacteria, yeast, and algae are capable of
accumulating metal ions extracellularly or
internally to concentrations several orders of
magnitude higher than the background
concentration, and many bacteria reduce
Cr(VI) to Cr(lll). Of the microorganisms studied,
S. cerevisiae was the only one that did not
result in an unpleasant or dangerous side
effect such as an unpleasant odor created in
the water or pathogenic results (Krauter et al.,
1996).
Microorganisms respond to metals by several
processes, including transport, biosorption to
cell biomass, entrapment in extracellular
capsules, precipitation, and oxidation-reduction
reactions. Bioaccumulation of metal cations
has been demonstrated by two processes: an
initial rapid accumulation that is independent of
metabolism and temperature, and a
metabolically mediated process that
internalizes the cation into the cell. Energy-
dependent uptake of divalent cations by S.
cerevisiae is well known, with influx being
dependent on the electrochemical proton
gradient across the plasma membrane (Krauter
etal., 1996).
Dissimilatory Metal-Reducing Bacteria:
Reduction of heavy metals, such as chromium,
through dissimilatory metal-reducing bacteria
(DMRB) has also been examined. DMRB gain
energy to support anaerobic growth by
coupling the oxidation of H2 or organic matter
to the reduction of a variety of multivalent
metals. This metabolism can lead to the
complete mineralization of organic matter or to
the precipitation and immobilization of metal
contaminants under anaerobic conditions. In
50
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situ bioremediation strategies using DMRB
would rely on either stimulating naturally
occurring DMRB populations or inoculating
preadapted or genetically engineered DMRB
into contaminated environments.
One difficulty of adding DMRB to a
contaminated area is that often vegetative
bacteria tend to bind to the substratum and are
rarely found far downstream from injection
wells. Starvation techniques have been
developed as a means of preparing bacteria
for in situ bioremediation. In order to survive
the oligotrophic conditions of certain natural
environments, many non-spore-forming
bacteria exhibit a starvation-survival response
under which cell size and susceptibility to
harsh conditions are reduced drastically.
Although these cells are metabolically dormant,
they are resuscitated to their vegetative state
when exposed to nutrients (Caccavo et al.,
1996).
Biomineralization: Biomineralization is the
microbially mediated genesis of new mineral
species. Specific microbial strains can be
employed to form geologically stable minerals
on solid substrates such as soils, fractures,
and ores. The microbial systems can be
engineered to form specific biominerals which
can incorporate and immobilize metal
contaminants (SAIC, 1998).
Biomineralization processes are part of a
natural cycle in which minerals are
continuously formed, transformed, and
degraded. In situ biomineralization capitalizes
on the role that microorganisms play in natural
ore formation and involves accelerating the
biological reactions to remediate waste.
Resarchers have evaluated the use of
bioremediation processes for in situ
biomineralization of heavy metals in mine
wastes. During biomineralization,
microorganisms initiate a complex series of
reactions. Effective metal removal mechanisms
are influenced by biologically catalyzed
remineralization reactions (Pintail Systems,
1998).
Biominerals are geologically stable compounds
that chemically bind the contaminants; other
polymeric and adsorption techniques do not
provide the same chemical stability.
Biominerals can be designed to be selective
for specific contaminants. Bioremediation fluids
can infiltrate pore spaces and micro fractures,
as opposed to cementatious compounds.
The biomineralization process can potentially
be applied to the subsurface treatment of
mobile metals by the formation of an in situ
biomineral barrier. The barrier is formed by
injecting bacterial and nutrient solutions into
the aquifer materials through a series of
inoculation (injection) wells. The biomineral
barrier functions as a reactive zone to specific
metal contaminants. The metal(s) are
incorporated into, and stabilized by, the
resultant biomineral product. The subsurface
barrier formation can be enhanced through
controlled hydro-fracturing.
Advantages
• Eliminates the infrastructure required for a
pump-and-treat system; no disposal of
water or waste.
• Inexpensive installation; primary capital
expenditure for this technology is the
installation of injection walls.
• Inexpensive operation; reagents are
injected at fairly low concentrations and
the only sampling required is groundwater
monitoring.
• Can be used to remediate deep site; no
physical limits as with treatment walls.
• Unobtrusive; once the system is installed,
site operations can continue without any
obstructions.
• Less expensive than most remediation
technologies.
• Immobilization of contaminant; uses the
capacity of the soils and sediments to
absorb, filter, and retain contaminants.
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• All biotic processes used for remediation
use simple sugars and bacteria natural to
the aquifer, therefore, it is completely
natural.
• DMRB starvation reduces cell size and
facilitates transport of bacteria through
substratum to the contaminated zone.
Limitations
• Longer time required for remediation
treatment.
• The metal is not actually removed from the
water, it is only put into a state in which
the water carrier will no longer interact with
it.
• Results in limited hydraulic control.
• Potential for short-circuiting or incomplete
treatment; barrier integrity verification is
more difficult.
• May not remove source of contamination;
mitigates contaminant plume.
• Low permeability sites may preclude use
of this method, but may be applicable to
source zone treatment.
3.3.2 Status
The in situ reactive zone approach is an
innovative and developing technology in the
remediation industry. Implementation and wide
acceptance of this technology is still in its
infancy, and thus the experience, knowledge,
and performance and cost data for this
technology is very much empirically based
(Suthersan, 1997). A substantial amount of
developmental work needs to be done on this
new technology before it reaches wide
regulatory acceptance. Future work should
focus on:
• Tools to design the appropriate
specification of injection rates, durations,
and concentrations to achieve optimal
control at the field scale.
• Tools to predict/estimate and measure the
target reaction kinetics in an in situ
environment.
• Tools to quantify reagent and pore water
chemistry at the field scale.
• Reactive transport modeling tools to
couple the microbial and chemical
reactions to the physical transport
processes.
• Better methods to measure the intra-
aqueous redox and biogeochemical
kinetics.
• Better understanding of the long-term fate
of the immobilized contaminants.
Laboratory and field testing has been
conducted with some of the biotic reactive
zone processes. These studies are described
in section 3.3.3.
3.3.3 Performance and Cost Data
Molasses-Based Reactive Zone
S/Ye 7: Afield pilot-scale demonstration test was
performed by investigators at an industrial
facility in the Midwestern United States to
evaluate this reactive zone remediation
technique involving the in situ reduction of
chromium. As of 1997, this evaluation involved
conducting a 6-month in situ test near the
source area at the site to determine the degree
to which Cr(VI) could be reduced and
precipitated out within the aquifer due to the
development of biologically induced reducing
conditions. The test was developed to evaluate
this innovative in situ remediation technique
that could potentially be used to augment or
replace the conventional pump-and-treat
system which was previously operated at the
facility. The field test required the installation of
three injection wells and five monitoring wells.
These wells added to the existing monitoring
well network at the facility.
52
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To promote the in situ biological reduction of
Cr(VI) to Cr(lll), a dilute water/blackstrap
molasses solution (200:1 dilution, by volume),
which contains readily degradable
carbohydrates and sulfur, was injected via
three injection wells (at a batch feed rate of
approximately 40 gallons every 2 weeks per
injection well) into the shallow portion of the
impacted aquifer. The installed injection and
monitoring wells were shallow wells screened
across a 1- to 3-ft thick sand seam at an
approximate interval of 10 to 15 ft below grade
(Suthersan, 1997).
Because of the rapid inducement of reducing
conditions, the concentration of Cr(VI) in the
injection wells decreased from a high of 15
mg/L to below 0.2 mg/L during the first month
of process operation. The levels of Cr(VI)
measured in the injection wells remained below
the cleanup objective of 0.2 mg/L through the
first 6 months of monitoring. The laboratory
analytical results for Cr(VI) measured in the
injection well samples collected 3 months
following process initiation were all below the
detection level of 0.05 mg/L, according to
investigators. In these same samples, the
levels of total chromium were slightly higher
than the levels of Cr(VI) but were still below the
0.2 mg/L groundwater cleanup objective for
this site. This indicates that chromium
remaining in the groundwater was in the
trivalent form, rather than in the more toxic and
mobile Cr(VI) form. In addition, it is important
to note that the analytical data were based on
unfiltered groundwater samples.
Because only trace amounts of Cr(lll) had
been detected in the unfiltered groundwater
samples, it appears that the chromium
precipitates were being retained by the aquifer
materials and were not subject to colloidal
transport through the aquifer (Suthersan,
1997).
Site 2: A full-scale application of the molasses-
based reactive zone was conducted at a site in
central Pennsylvania. Groundwater was
impacted by Cr(VI) downgradient of an
operating manufacturing facility. This site was
placed on the NPL list in 1988 and a ROD was
issued in 1991 by EPA. Following a successful
pilot-study demonstration, a full-scale
remediation system was installed to develop
and maintain an anaerobic environment
capable of reducing and precipitating Cr(VI).
Figure 3-11 shows the molasses injection
system that was installed at the facility and that
went on line in January 1997. The system
utilized 20 installed injection wells and 16
existing municipal wells to establish reactive
zones. Ten gallons of solution per well were
injected twice a day. The mixing ratio for
molasses was varied from 1:200 to 1:20. A
programmable logic controller (PLC) monitored
and controlled the feedrate and frequency of
the molasses feed and solution feed pumps,
as well as the timing of the solenoid valve
network that controlled the metered flow to the
injection wells. Monthly to quarterly sampling
has been conducted for pH, redox levels, and
chromium concentrations (Burdick and Jacobs,
1998).
53
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Molasses Injection System
Feed Pump Potable Water
Reactive Zone
Building
Source: Lenzo, 1999
Piping Network Injection Wells
Figure 3-11. Schematic of molasses-based injection system used at central Pennsylvania
site.
The injection of the molasses-based reagent
was successful in creating a anaerobic
groundwater environment and has resulted in
Cr(VI) concentrations being decreased
significantly. Figures 3-12 and 3-13 show how
the extent and concentrations of Cr(VI)
groundwater plume were reduced in a year and
a half. The overall chromium plume shrunk to
approximately one-fourth its original area. The
concentration of Cr(VI) was reduced from 1.95
mg/L to 0.01 mg/L in the southern portion of
the treatment area. The peak chromium
concentrations are isolated to one area at
slightly above 0.5 mg/L, as shown in Figure
3-13.
After close to 3 years of operation, chromium
concentrations were reduced to below the
regulatory target across the entire site,
according to investigators (Lenzo, 1999).
The cost to implement the molasses-based
reactive zone technology and operate it for a
little less than 3 years was approximately
$400,000 including capital, O&M, and
monitoring. This system replaced a pump-and-
treat system that had an estimated present
worth of over $4,000,000. This figure included
capital costs and O&M for a period of 20 years
(Lenzo, 1999).
54
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LEGEND
* Injection Well
' Monitoring Well
o.s— Cr (VI)
Treatment Building
Source: Lenzo, 1999
Figure 3-12. Plot of hexavalent chromium contaminant plume at central Pennsylvania site
- January 1997.
LEGEND
* Injection Well
- Monitoring Well
o.i— Cr (¥1)
Source; Lenzo, 1999
Figure 3-13. Plot of hexavalent chromium contaminant plume at central Pennsylvania site
-July 1998.
55
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Other Biotic-Based Reactive Zones
Unicellular Yeast: Laboratory studies were
performed using amended groundwater to
investigate the use of the yeast S. cerevisiae as
an agent to remove Cr(VI). These studies also
examined the effects of pH, temperature, and
energy source concentration on Cr(VI) removal.
Results of these studies showed S. cerevisiae
removed Cr(VI) at the moderate rate of 0.227
mg/h (g dry wt biomass)"1 (Krauter et al., 1996).
DMRB: Experiments were conducted exam-
ining the effects of starvation on a model
DMRB, Shewanella alga BrY. Shewanella alga
bacteria were shown to reduce Fe(lll) to Fe(ll),
which in turn reduces Cr(VI). By starving the S.
alga BrY, thereby reducing its cell size and
endogenous metabolic activity, and
resuscitating it with a variety of electron
acceptors, including oxygen, Fe(lll) and natural
subsurface materials, the DMRB can be
delivered into a reactive zone contaminated with
chromium in a faster, more effective manner
(Caccavo, 1996). Studies have shown that
starved S. alga BrY can reduce 90 percent of
Fe(lll) in subsurface material to Fe(ll) within 4
days (Caccavo, 1996).
Biomineralization: Work to date with biomin-
eralization has focused on the treatment of
metal bearing ores and mine process solutions.
Treatment strategies have been applied both in
situ as well as ex situ. This technology was
accepted into the EPA Emerging Technology
Program in 1995 for evaluation. Further
development of the process could result in a
field-ready in situ biomineralization technology.
However, biomineralization is still an emerging
technology, and is not yet proven for in situ
applications. Investigators have demonstrated
biomineralization of metals in laboratory and
pilot-scale tests for mining industry clients at
mines in the U.S. (Idaho, Nevada, Arizona,
California, Colorado), Mexico, and Canada.
Performance and cost data for in situ metals
treatment using the biomineralization process
were not available during preparation of this
Guide.
3.4 Soil Flushing/Chromium
Extraction
3.4.1 Technology Description
In situ soil flushing is used to mobilize metals
by leaching contaminants from soils so that
they can be extracted without excavating the
contaminated materials. Water or an aqueous
solution is injected into or applied onto the area
of contamination to mobilize the contaminants.
The flushing solution can be applied by surface
flooding, sprinklers, leach fields, vertical or
horizontal injection wells, basin infiltration
systems, or trench infiltration systems. After
contact with the contaminated material, the
flushing solution is collected using pump-and-
treat methods for disposal or treatment and
reuse.
In situ soil flushing can enhance conventional
pump-and-treat by providing a hydraulic push
in an aquifer, by increasing the hydraulic
gradient, and by solubilizing/mobilizing metal
contaminants more rapidly. This can result in
an accelerated rate of contaminant removal
(Steimle, 1997).
Metal contaminants are mobilized in situ by
solubilization, formation of emulsions, or a
chemical reaction with the flushing solution.
Various water treatment techniques can be
applied to remove the extracted metals and to
recover the extraction fluid (if other than water)
for reuse. The separation of surfactants from
recovered flushing fluid for reuse in the
process is a major factor in the cost of soil
flushing. Treatment of flushing fluid results in
process sludges and residual solids. Residual
flushing additives in the soil may be a concern
and should be evaluated on a site-specific
basis. Figure 3-14 shows a generalized
schematic of the in situ flushing process.
56
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•Si-
Derived frame Steimle, 1997
Figure 3-14. Schematic of in situ flushing system.
Subsurface containment barriers (e.g., a slurry
wall) can be used in conjunction with soil
flushing technology to help control the flow of
flushing fluids. Soil flushing is most applicable
to contaminants that are relatively soluble in
the extracting solution, and that will not tend to
sorb onto soil as the metal-laden flushing fluid
proceeds through the soil to the extraction
point. The selected flushing fluid must be
compatible with the metal(s) of concern and
the soil properties to work effectively. Optimal
conditions for extraction of Cr(VI) using soil
flushing are permeable soil with low iron oxide,
low clay, and high pH. Flushing enhancements
that are currently being researched include
acids, bases, chelating agents, and
surfactants/cosolvents to aid in the
desorption/dissolution of the target metals
(U.S. EPA, 1997). The use of chelating
additives for treating metals in situ has not yet
been found to be effective (USEPA, 1996).
Surfactant-enhanced extraction can be used to
expedite the removal of chromium from source
zone soils in order to mitigate the continual
leaching of the contaminant into the
groundwater plume. Lately, surfactant-
enhanced pump-and-treat remediation has
received considerable attention. Surfactants
have an amphiphilic structure that results in
their surface active nature, and causes them to
concentrate in interfacial regions (Sabatini et
al., 1997). A surfactant adsorbs to interfaces
and significantly decreases the interfacial
tension and alters the wetting properties of the
soil matrix (Palmer and Fish, 1992). It is
hypothesized that surfactants can displace
adsorbed chromate by either ion exchange,
precipitation-dissolution, and/or counterion
57
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binding mechanisms, and that further
enhancement in extraction may be achieved if
surfactants with solubilized complexing agents
are used (Sabatini et al., 1997). In effect, the
use of surfactants helps to overcome mass
transfer limitations for contaminant removal.
Soil flushing can be applied to source zone
soils if additives are used to accelerate
treatment. Soil flushing with water alone can
be applied to the concentrated or dilute
portions of the groundwater plume, depending
on how easily the metals contamination can be
mobilized.
The applicability of in situ soil flushing
technologies to contaminated sites will depend
largely on site-specific properties, such as
hydraulic conductivity, that influence the ability
to contact the contaminants with the extractant
and to effectively recover the flushing solution
with collection wells (Evanko and Dzombak,
1997).
Some factors that are critical for the success of
soil washing for chromium and metals
extraction include hydraulic conductivity, cation
exchange capacity (CEC), clay content, and
carbon content. Sites with higher hydraulic
conductivity are more conducive to effective
soil flushing. Higher levels of CEC, clay
content, and organic carbon increase sorption
and inhibit metals contaminant removal.
Aboveground sprayers, infiltration galleries,
and injection/extraction wells can be used to
accomplish flushing of soil and groundwater.
Some equipment can be mobile. Tanks or
ponds are needed for washwater preparation
and wastewater treatment. Slurry walls or other
containment structures may be needed along
with hydraulic controls to ensure capture of
contaminants and flushing additives. Cold
weather freezing must be considered for
shallow infiltration galleries, aboveground
sprayers, and extracted wastewater treatment.
Permits may be required for operation, air
discharges, and injection of flushing additives,
depending on the system being utilized and
the contaminants of concern (U.S. EPA, 1997).
Advantages
• Can potentially accelerate removal of
chromium in source areas and meet clean-
up goals.
• Removal of chromium source
contamination will beneficially impact
down gradient groundwater plume.
• Contaminant is removed from the soil or
aquifer material and may be applicable to
recovery.
• Eliminates the need to excavate, handle,
and dispose of large quantities of
contaminated soil.
• Can reuse some flushing solutions after
treatment for separation, providing cost
savings.
• Equipment used for the technology is
relatively easy to construct and operate.
• Useful in treatment train applications.
Limitations
• May be more applicable to organic
contaminants than metals.
• Still in the developing stage; limited field
experience.
• Not considered Resource Conservation
and Recovery Act (RCRA) Best
Demonstrated Available Technology
(BOAT) for chromium, lead, mercury,
arsenic, and cadmium.
• May not be applicable for certain site
characteristics such as low hydraulic
conductivity and high organic matter
content.
• May be difficult to apply to sites
contaminated with more than one type of
metal.
58
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• Treatment time may be very long unless
additives are used for enhanced
treatment.
• Surfactant solutions may leave residual
surfactant in aquifer materials.
• Has a potential for spreading
contaminants horizontally or vertically.
• It may be difficult to gain regulatory
approval for injection of surfactant
solutions due to concerns over residuals
and toxicity.
3.4.2 Status
In the U.S., where full-scale site remedies have
utilized in situ flushing, water is typically used
as the flushing solution. This technology has
been applied for a limited number of metal-
contaminated sites; it is still in a developing
stage. There has been more research and
application of the technology for organic forms
of contamination. At least two applications of
soil flushing with water for chromium removal
have been documented: the United Chrome
Products Superfund site in Corvallis, Oregon,
and the Lipari Landfill site in New Jersey.
Remediation at the United Chrome site began
in 1985 and has used infiltration basins and
trenches to flush contaminated soils, a 23-well
groundwater extraction network, on-site
treatment of wastewater, and off-site disposal
of contaminated soil and debris (Evanko and
Dzombak, 1997; USEPA, 1997).
Sabatini et al. (1997) and Nivas et al. (1996)
showed the ability of surfactants to enhance
the elution of chromate in column studies
significantly. Their laboratory batch and column
studies also evaluated hypotheses concerning
the displacement of adsorbed chromium by
surfactants and complexing agents. Their work
showed that while this technology has great
promise for enhancing chromium extraction
from soils, further laboratory and field-scale
studies are necessary to evaluate operational
considerations prior to full-scale
implementation. This work was conducted
using chromium-contaminated soil samples
from the USCG Support Center site in
Elizabeth City, North Carolina.
Researchers are also investigating the effects
of numerous soil factors on heavy metal
sorption and migration in the subsurface. Such
factors include pH, soil type, soil horizon, CEC,
particle size, permeability, specific metal type
and concentration, and type and
concentrations of organic and inorganic
compounds in solutions (U.S. EPA, 1997).
Major concerns for in situ flushing are the
uncertainty of the fate and effects of washing
reagents in the subsurface environment, and
preventing mobilized contaminants from
migrating into the surrounding environment
(Yin and Allen, 1999).
3.4.3 Performance and Cost Data
United Chrome Products Superfund Site
At the United Chrome Products site, soil and
groundwater were heavily contaminated with
chromium, having total chromium levels in the
soil as high as 60,000 mg/kg and levels in the
groundwater reaching up to 19,000 mg/L. The
in situ flushing procedure used at this site
leached contaminants from the unsaturated
and saturated zones, and provided for
recharge of the groundwater to the extraction
wells. According to investigators, this cleanup
operation removed significant amounts of
chromium from the soil and groundwater, and
the pumping strategy achieved hydraulic
containment of the plume. Cr(VI) levels in
extracted groundwater decreased from a
maximum measured concentration of more
than 5,000 mg/L to approximately 50 mg/L
during the first 2.5 years of operation using
water flushing. The average chromium
concentration from multiple measurements in
the groundwater plume decreased from 1,923
mg/L to 207 mg/L after flushing the first 1.5
pore volumes (approximately 2.6 million
gallons for one pore volume). This removal rate
was expected to continue for the first few pore
volumes of treatment until Cr(VI) removal
began to tail off to the asymptotic level
(Sturgesetal., 1992).
59
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Elizabeth City, North Carolina USCG Site
Batch and column extraction studies using
surfactants were performed on chromium-
contaminated source zone soils from the
Elizabeth City, North Carolina USCG site.
These studies were designed to evaluate the
efficiency of chromate extractions for selected
surfactants and complexing agents, and to
evaluate the mechanisms of removal. In the
batch studies, the amount of Cr(VI) removed
was observed to increase with anionic
surfactant concentration. At concentrations
greater than the critical micelle concentration
(CMC), where micelles or droplets form, Cr(VI)
removal was relatively constant.
enhanced the extraction of Cr(VI) by an order
of magnitude greater than that obtained with Dl
water.
The column study results supported the batch
study results. As seen in Table 3-6, the ratio of
Cr(VI) removed by surfactant (with or without
the solubilizing complexing agent, diphenyl
carbazide (DPC)) was greater than for Dl water
only, and the number of pore volume flushes
required was less with the surfactants. These
studies demonstrated that surfactant-enhanced
systems have the potential to enhance
chromium extraction by a factor of 2 to 3
versus water alone, and surfactant with
complexing agent systems can enhance
Table 3-5. Results of Cr(VI) Extraction Studies by Surfactants
and Hydrotropes from Elizabeth City Soil.
Extracting
Agent
D.I. Water
AOT
SDS
Dowfax8390
Deriphat-1 60
Extracting Agent Cone, at
Max. Cr(VI) Removal, mM
—
3
10
1
5
Ratio to CMC
—
2.7
1.2
0.3
—
Max. Cr(VI)
Cone., ppm
2.6
5.2
6.4
5.6
7.4
Ratio of Cr(VI) Removal
by Extracting Agent
to That by Water
1
2.0
2.5
2.1
2.8
Source: Sabatini et al., 1997
The ratio of maximum Cr(VI) removal by
surfactants to that of deionized (Dl) water
ranged from 2.1 for the surfactant Dowfax
8390, to 2.8 for the surfactant Deriphat-160
(see Table 3-5). It was postulated, based on
test results, that ion exchange was the primary
extraction mechanism. Tests were also
conducted to see if solubilizing complexing
agents would enhance Cr(VI) extraction. In all
cases, the surfactant with solubilized
complexing agent additive outperformed the
surfactant only results; the addition of a
chromium (solubilizing) complexing agent
chromium extraction by an order of magnitude
greater versus water alone systems. The
researchers also concluded that operational
considerations for surfactant extraction of
chromium require additional research. For
example, it was suggested that the removal of
Cr(VI) from the soil can be further enhanced by
optimizing the time of switching from injection
of surfactant with DPC to surfactant alone
and/or by increasing the surfactant
concentration. Laboratory and field studies are
critical prior to full-scale implementation
(Sabatini etal., 1997).
60
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Table 3-6. Cr(VI) Extraction From Columns by Water, Surfactants
Alone, and Surfactant Solubilized DPC
Extracting
Agent
Dl Water
ACT
Dowfax 8390
ACT with DPC
Dowfax with DPC
Max. Cr(VI)
Removed in
Effluent, ppm
5.3
7.0
11.8
10.3
19.7
Ratio of Cr(VI)
Removed by
Extracting Agent
to that by D.I. Water
1
1.3
2.2
1.9
3.7
35.9
26.5
24.4
24.3
18.9
Source: Sabatini et al., 1997
Cost data for chromium or metals-specific site
remediations using in situ soil
flushing/extraction were not available during
preparation of this Guide. However, cost
estimates for use of the technology on a variety
of contaminants have been compiled from EPA
documents. One source provides an estimated
cost range for soil flushing of $60 to 163/ton
(U.S. EPA, 1997). These cost estimates
generally do not include pretreatment, site
preparation, regulatory compliance costs, costs
for additional treatment of process residuals, or
profit. The actual cost of employing soil flushing
technology at a specific site may be
significantly different than these estimates.
Another source estimates the operating costs
for soil flushing/extraction technology at
approximately $70 to 170/ton (Evanko and
Dzombak, 1997). The initial and target
contaminant concentrations, soil permeability,
and the depth of the aquifer will influence
costs. Chemically enhanced (surfactants, etc.)
flushing systems will have additional costs
associated with reagents and equipment
needed to handle the flushing solution.
3.5 Electrokinetics
3.5.1 Technology Description
The theory of applying electric current to
groundwater for remediation of heavy metal
and other wastes is called electrokinetic
remediation. It is also called electroreclamation
and electrochemical decontamination.
Electrokinetics is a process that separates and
extracts heavy metals, radionuclides, and
organic contaminants from saturated or
unsaturated soils, sludges, and sediments.
Basically, a series of electrodes are placed in a
contaminated area to which a low voltage (50
to 150 volts) direct charge is then applied.
Because of the charge on water and the
contaminant, desorption and subsurface
migration will occur towards the oppositely
charged electrodes. Active electrodes in water
cause an acid front at the anode and a base
front at the cathode. The pH will drop at the
anode and increase at the cathode. To prevent
this pH imbalance, the electrodes are placed
inside ceramic casings which are filled with a
processing fluid. This processing fluid not only
keeps a balance of pH at the anode and the
cathode, it can also help solubilize and move
contaminants. Electrokinetic treatment
concentrates contaminants in the solution
around the electrodes. Contaminants can be
removed from this solution by electroplating or
precipitation/coprecipitation at the electrodes,
or by pumping the contaminant and processing
fluid to the surface and treating with ion
exchange resins other methods to recover the
extracted metal and reuse the processing fluid
in the electrokinetic system. Figure 3-15 is a
schematic of the process.
61
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Extraction/
Exchange
II
Processing
Process Control System
r
r ~" ~"
i
i
Extraction/
Exchange
II
1 * ~- — ~- -~ I
i Processing
1
- CATHODE
~ Base Front
and/or Cathodic
Process Flluid
Front
Anodic
Fluid
Figure 3-15 Electrokinetic remediation process.
Electrokinetic remediation is possible in
saturated and unsaturated soils. The soil
moisture content must be high enough to allow
electromigration, but for optimum results, should
be less than saturation, to avoid the competing
effects of tortuosity and pore water content.
Surfactants and complexing agents can be used
in subsurface media to increase solubility and
assist in the movement of the contaminants.
Also, reagents may be introduced at the
electrodes to enhance contaminant removal
rates. The efficiency of metal removal by this
process will be influenced by the type and
concentration of contaminant, the type of soil,
soil structure, and interfacial chemistry of the soil
(U.S. EPA, 1997b and 1995a; Evanko and
Dzombak, 1997). Conditions that can produce
optimal electrokinetic performance include soil
moisture near saturation, adequate pore water
electroconductivity, metal solubility, low CEC,
and low salinity (Van Cauwenberghe, 1997).
Equipment required for the in situ field
application of this technology is of a fairly
specialized nature. Anodes and cathodes are
placed in permeable or porous casings in situ.
The aboveground system requires a pump to
remove contaminated water from the cathode
to a processing system. Tanks and meters are
needed for holding waste to be processed, and
water solutions or chemical additives that are
used in situ. A low voltage power supply is
required. Other specialized equipment such as
controllers, valves, vacuum pumps, and
gauges may be required.
In situ electrokinetic remediation is primarily for
the remediation of sites with low permeability
and in order to overcome subsurface
heterogeneities. Several variations of the
electrokinetic process include:
• A technology called the "Pool Process" is
used to remediate toxic heavy metals such
as chromium. Through this process, ion-
permeable electrolyte casings are placed in
the contaminated media and connected to a
centralized electrochemical ion-exchange
(EIX) based electrolyte management
system.
62
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Each casing has an electrode inside.
Together, these form alternating rows of
anodes and cathodes. Electrolyte is
circulated in a closed loop between the
electrode casings and the EIX. Electrolysis
of water in the electrolyte results in the
formation of H+ ions at the anodes and OH-
at the cathodes. These ions migrate
through the casing into the soil generating
a temporary and localized pH shift that
desorbs contaminating ions. Once
desorbed, the contaminating ions migrate
under the influence of the applied potential
to their respective electrodes (anodes for
anions, cathodes for cations). Here they
pass through the electrode casing walls and
are taken up by the circulating electrolytes.
The pH at the anode and the cathode is
managed by the addition of acid or alkali,
as required. Contamination is selectively
recovered from the circulating electrolytes
as they pass through the EIX units. Soluble
but benign elements are returned to the
media. Periodically, the EIX units are
regenerated by polarity reversal, which
recovers the contamination in a
concentrated and reusable form (U.S. EPA,
1997b).
Another process operates similarly to the
process described previously. It is
potentially applicable to saturated and
unsaturated soils. Conditioning pore fluids
may need to be added to the soil matrix or
circulated at the electrodes to control
process electrochemistry. The process is
being used to stimulate and sustain in situ
bioremediation for treatment of organics
and heavy metals. This is done by
introducing nutrients and process additives
to the subsurface.
Theoretically, the rate of additive transport
and the efficiency of its dispersion in the
subsurface matrix will be enhanced by the
use of electrokinetics, especially at sites
with aquifer heterogeneities (U.S. EPA,
1996a).
• An in situ electrokinetic extraction (ISEE)
system is being developed that can be
used to treat anionic heavy metals such as
chromate in unsaturated soil without
adding significant amounts of water to
control process electrochemistry. Water is
only circulated (added) to the electrode
casing to help remove collected
contaminants. Bench-scale studies have
shown the technology to be effective in
sandy soils with a moisture content as low
as 7 percent. The technology can be
expanded to treat saturated soils (U.S.
EPA, 1996a).
Some of the potential advantages of the use of
electrokinetic technology are listed below.
Advantages
• Effective method for inducing movement
of water and ions through fine-grained,
low-permeability, or heterogeneous soils
that may be an obstacle to more traditional
technologies.
• Mobilizes metal contaminants without use
of strong acids for pH modification.
• May be used to remediate heavy metals
contamination in unsaturated soils;
technically and cost competitive.
• Applicable to a broad range of
contaminants.
A number of potential limitations in using this
technology exist; most of these limitiations are
site-specific. These limitations are listed below
(Van Cauwenberghe, 1997):
Limitations
• The contaminant needs to be solubilized
either by a dilute acid solution front or by
a processing fluid in order for it to be
extracted by most electrokinetic
processes.
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• Process is limited by the solubility of the
contaminant and the desorption of
contaminants from the soil matrix.
• Process may not be efficient for treating
multiple metals if concentrations are
significantly different.
• Incomplete remediation could result if there
are areas of poor electrical conductivity
(stagnant zones) between wells, or the
contaminant migration path is long.
• Heterogeneties or subsurface anomalies
such as building foundations or large rocks
can reduce removal efficiencies.
• Immobilization of metal ions can occur by
undesirable chemical reactions with
naturally occurring and co-disposed
chemical constituents.
• Heavy metals can prematurely precipitate
close to the cathode.
• Excessive treatment depths may not be
cost-effective for use of the technology.
3.5.2 Status
Electrokinetic remediation is a developing
innovative technology with a specialized nature.
Its main focus has been the treatment of low
permeability soils where other technologies
would not be successful or their use may not be
cost-effective. Because of its specialized nature,
relatively few commercial vendors apply the
technology (Van Cauwenberghe, 1997).
The success of various electrokinetic
remediation technologies for removal of
metals, including chromium, from soils has
been shown via bench- and pilot-scale
experiments. Currently, several of these
technologies are being implemented in
comprehensive field-scale demonstration
studies for further evaluation (Evanko and
Dzombak, 1997).
A technology vendor using the Pool Process
continues to perform several electrokinetics
remediation projects both in the U.S. and
abroad (see Table 3-7). The commercial scale
electrokinetic remediation technology is mainly
used for the extraction of toxic metals and toxic
anions from soil and groundwater (U.S. EPA,
1997b). The technology has also been
demonstrated under the Superfund Innovative
Evaluation Technology (SITE) Program in
chromate-contaminated soil atSandia National
Laboratories (SNL), in New Mexico (U.S. EPA,
1996a).
Other processes have also been demonstrated
under the SITE Program. A field pilot-scale test
was conducted at a site contaminated with
lead, copper, and zinc. A field study of the
ISEE Process was conducted at an unlined
chromic acid pit within a landfill (U.S. EPA,
1996a).
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Table 3-7. Case Studies for the Pool Process
Year
1997- ongoing
1996
1992-1994
Location
Alameda Naval
Air Station,
California
California
Temporary landfill
at the Airbase of
Woensdrecht,
Germany
Client
U.S. Navy and EPA
Office of Tech.
Development
Large U.S. industrial
and communications
company
Ministry of
Defense/DGWT
Description
Pilot-scale recovery of Cr from
former plating operations
Bench scale recovery of Cr
from former plating operations
Formed on-site lagoon and in
situ remediation of 3,400 yd3
sludge contaminated with Cr,
Ni, Cu, Zn, and Cd
ongoing
$16,000
$1,040,000
3.5.3 Performance and Cost Data
Most of the applications of this technology to
date have been for bench- or pilot-scale testing.
There has been very little actual in situ
remediation of chromium at full-scale using this
technology. Electrokinetic remediation of metals
in situ is still a developing technology. Due to
the developing nature of the technology,
available performance and cost data are
estimated and should be used with caution.
The Pool Process has had the most application,
mostly at sites in Europe, and often as an ex
situ treatment or on sediment lagoons, not on
undisturbed sites where some of the listed
advantages of the technology can be evaluated.
Investigators estimate that concentrations of
target species in the range of 10 to 500 ppm
can be reduced to less than 1 ppm.
Remediation costs are expected to be in the
range of $200 to $325/m3 ($150 to $250/yd3)
(Van Cauwenberghe, 1997).
Testing of another electrokinetics process has
shown removal efficiencies between 75 and 95
percent for lead, chromium, cadmium, and
uranium at levels up to 2,000 mg/kg (Van
Cauwenberghe, 1997). Bench-scale testing at
SDL in sandy soils at approximately 40 to 60
percent soil moisture saturation resulted in
removal by the process of 75 to 90 percent of
the initial chromium (USEPA, 1995).
A field-scale demonstration of the ISEE System
for treatment of chromate-contaminated soil
was conducted at SDL under the SITE Program
(USEPA, 1998). The ISEE System was
developed to remove Cr(VI) from unsaturated
soil. The field-scale demonstration results
showed that the ISEE removed approximately
200 g of Cr(VI) during operation, and had an
overall removal efficiency of approximately 0.14
g of Cr(VI) per kilowatt hour (kWh). However,
comparison of pre- and post-treatment soil
sample results did not show much improvement
in Cr(VI) levels or TCLP levels in the treatment
zone. The post-treatment median TCLP
concentration of 20.4 mg/L exceeded the TCLP
regulatory limit of 5.0 mg/L for the
demonstration site. The total treatment costs for
the ISEE System to treat 16 yd3 of soil were
estimated to be $1,830/m3 ($1,400/yd3) for
removing 200 g of Cr(VI). This cost will vary
depending on cleanup goals, soil type,
treatment volume, and system design changes.
The ISEE System used for the SITE
demonstration was a prototype. The treatment
cost for a full-scale system should be
significantly reduced due to design
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improvements based on the SITE
demonstration results (U.S. EPA, 1998).
The cost of using electro kinetic remediation is
dependent on specific chemical and hydraulic
properties present at the contaminated site.
Cost is strongly influenced by soil conductivity
because energy consumption is directly related
to the conductivity of the soil between the
electrodes. Electrokinetic treatment of soils with
high electrical conductivities may not be feasible
due to the high cost (Evanko and Dzombak,
1997). Estimated price ranges per unit of waste
(non-specific) treated by various vendors
include:
• DuPont R&D: $85/m3 ($65/yd3)
Electrokinetics, Inc.: $25 to $130/m3 ($20
to$100/yd3)
Geokinetics International: $80 to $300/m3
($60 to $225/yd3)
These price estimates do not include indirect
costs associated with remediation, such as
permits and treatment of residues (Van
Cauwenberghe, 1997).
Other factors that have a significant effect on
unit price are:
Depth of contamination: greater treatment
depth interval is more cost effective.
Residual waste processing.
Site preparation and system installation
requirements.
Local electricity and labor costs.
Pilot-scale field studies indicate that the energy
consumption in extracting heavy metals from
soil may be approximately 500 kW-hr/m3 or
more at an electrode spacing of 1.0 to 1.5 m.
The direct energy cost would be approximately
$25/m3 ($20/yd3) or $0.05/kW-hr at this level of
energy consumption.
3.6 Natural Attenuation
3.6.1 Technology Description
The EPA accepted form of natural attenuation,
"monitored natural attenuation," refers to the
reliance on natural attenuation processes
(within the context of a carefully controlled and
monitored site cleanup approach) to achieve
site-specific remedial objectives within a time
frame that is reasonable compared to that
offered by other more active methods. The
natural attenuation processes that are at work
in such a remediation approach include a
variety of physical, chemical, or biological
processes that, under favorable conditions, act
without human intervention to reduce the mass,
toxicity, mobility, volume, or concentration of
contaminants in groundwater. These in situ
processes include biodegradation; dispersion;
sorption; dilution; volatilization; and chemical or
biological stabilization, transformation, or
destruction of contaminants. Other commonly
used terms referring to natural attenuation
include "intrinsic remediation," "intrinsic
bioremediation," "passive bioremediation,"
"natural recovery" and "natural assimilation"
(U.S. EPA, 1997d).
Use of Natural Attenuation
The use of monitored attenuation is not new to
remediation. It has been incorporated in
contaminant remediation plans since 1985.
Since that time, monitored natural attenuation
has continued, slowly increasing with greater
program experience and scientific
understanding of the processes involved.
Though recent scientific advances have
resulted in a heightened interest in this
approach, complete reliance is appropriate only
in a limited set of circumstances at
contaminated sites (U.S. EPA, 1997d).
Natural attenuation processes are typically
occurring at all sites, but to varying degrees of
effectiveness depending on the types and
concentrations of contaminants present and the
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physical, chemical, and biological
characteristics of the soil and groundwater.
Natural attenuation processes may reduce the
potential risk posed by site contaminants in a
number of ways: the contaminant may be
converted to a less toxic form through
destructive processes such as biodegradation
or abiotic transformations; potential exposure
levels may be reduced by lowering of
concentration levels by dilution or dispersion;
contaminant mobility and bioavailability may be
reduced by sorption to the soil or rock matrix;
Cr(VI) may be reduced to the less toxic and
mobile Cr(lll) by natural reductants; and metal
contaminants may be incorporated into the
crystalline structure of a rock or mineral.
Following source control measures, natural
attenuation may be sufficiently effective to
achieve remediation objectives at some sites
without the aid of other (active) remedial
measures. Typically, however, natural
attenuation will be used in conjunction with
active remediation measures (U.S. EPA,
1997d). Natural attenuation is often applicable
to sites that are not highly contaminated.
Usually the source area has been treated
through active remediation measures that are
then followed with natural attenuation to
complete the remediation of residual
contamination.
Sorption and oxidation-reduction (redox)
reactions are the dominant mechanisms
responsible for the reduction of mobility, toxicity,
or bioavailability of inorganic contaminants. It is
necessary to know what specific mechanism is
responsible for the attenuation of inorganic
contaminants because some mechanisms are
more desirable than others. In the case of
chromium, it is critical to have an understanding
of the Chromium Cycle in the environment.
Changes in a contaminant's concentration, pH,
redox potential, and chemical speciation may
reduce a contaminant's stability at a site and
release it into the environment. Determining the
existence and demonstrating the irreversibility
of these mechanisms are key components of a
sufficiently protective monitored natural
attenuation remedy (USEPA, 1997d).
Site Characterization
If natural attenuation is to be considered as a
viable option for chromium contaminated sites,
then ideally, it must be demonstrated that: (1)
there are natural reductants present within the
aquifer; (2) the amount of Cr(VI) and other
reactive constituents does not exceed the
capacity of the aquifer to reduce them; (3) the
time scale required to achieve the reduction of
Cr(VI) to the target concentration is less than
the time scale for the transport of the aqueous
Cr(VI) from the source area to the point of
compliance; (4) the Cr(lll) will remain immobile;
and (5) there is no net oxidation of Cr (III) to
Cr(VI). The most difficult information to obtain is
the time scale for the reduction and oxidation of
chromium in the soil (Palmer and Puls, 1994).
Decisions to employ monitored natural
attenuation as a contaminant remedy or remedy
component should be thoroughly and
adequately supported with site-specific
characterization data and analysis. In general,
the level of site characterization necessary to
support a comprehensive evaluation of natural
attenuation is more detailed than that needed to
support active remediation. For example, to
assess the contributions of sorption, dilution,
and dispersion to natural attenuation of
chromium-contaminated groundwater requires
a very detailed understanding of aquifer
hydraulics, recharge and discharge areas and
volumes, and chemical properties of the aquifer
system (U.S. EPA, 1997d).
Once the site characterization data have been
collected and a conceptual model developed,
the next step is to evaluate the efficacy of
monitored natural attenuation as a remedial
approach. This approach would more likely be
appropriate if the plume is not expanding or
threatening downgradient wells or surface water
bodies, and where ample potable water
supplies are available.
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Monitoring
Performance monitoring is of even greater
importance for monitored natural attenuation
than for other types of remedies due to the
longer remediation time frames, potential for
ongoing contaminant migration, and other
uncertainties associated with it. Monitoring
programs developed for each site should
specify the location, sampling frequency, and
type of samples and measurements necessary
to evaluate technology performance as well as
define the anticipated performance objectives of
the technology. Performance monitoring should
continue as long as contamination remains
above required cleanup levels. Additionally,
monitoring is continued for a specified period
after cleanup levels have been achieved to
ensure that concentration levels are stable and
remain below target levels (U.S. EPA, 1997d).
Demonstrating Cr(VI) reduction in the aquifer by
mass balances that rely primarily on the
aqueous concentrations from monitoring well
networks is valid only if it is demonstrated that
Cr(VI) precipitates are not forming in the
aquifer. The monitoring network must be
sufficiently dense to ensure that estimates of
Cr(VI) are accurate (Palmer and Puls, 1994).
Contingency Remedy Technologies
It is often suggested that a facility using natural
attenuation as a remediation remedy have a
contingency remedy established. A contingency
remedy is a cleanup technology or approach
specified in the site remedy decision document
that functions as a "backup" remediation
approach in the event that the "selected"
remedy fails to perform as anticipated. It is also
recommended that one or more criteria
("triggers") be established, as appropriate, in the
remedy decision document that will signal
unacceptable performance of the selected
remedy and indicate when to implement
contingency measures (U.S. EPA, 1997d).
As chromium concentrations decrease while
using conventional, active remedial procedures,
it becomes more difficult to remove the
remaining chromium. These active procedures
can also be very costly. In response to these
obstacles, natural reductants have been
identified that transform the more toxic
hexavalent form of chromium to the less toxic
trivalent form. Under alkaline to slightly acidic
conditions (pH>4.0), Cr(lll) precipitates as a
fairly insoluble hydroxide, thereby immobilizing
it (Bartlett and Kimball, 1976a). Such natural
attenuation may mean that strict water-quality
standards do not have to be attained
everywhere within and beneath the site. For
instance, conventional pump-and-treat
remediation could desist after the most
contaminated groundwater has been removed,
even if the MCL has not been achieved, if it is
anticipated through analysis that natural
attenuation will deal with the residual
contamination. Under certain circumstances,
expensive remedial measures may not even be
necessary (Palmer and Puls, 1994).
There are several soil tests that are useful in
determining the mass of Cr(VI) and Cr(lll) in the
source areas and the reduction and oxidation
capacities of the aquifer materials. Using
conceptual models, this information, combined
with knowledge of the residence time of the
chromium between the source and the point of
compliance, can be used to determine the
feasibility of natural attenuation of Cr(VI). The
major limitation to this approach is the lack of
information about the rate of oxidation and
reduction of chromium under conditions likely to
be encountered by plumes emanating from
chromium sources. Until better information is
developed about these rate processes under a
wider range of conditions with respect to pH,
the use of natural attenuation for contaminated
soils and groundwater will continue to be a
highly debated issue (Palmer and Puls, 1994).
Advantages
• Much less costly than other in situ remedial
technologies.
• Can remove the low concentration levels of
chromium that pump-and-treat systems are
unable to remediate.
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• Generation of lesser volume of remediation
wastes, reduced potential for cross-media
transfer of contaminants commonly
associated with ex situ treatment, and
reduced risk of human exposure to
contaminated media.
• Less intrusion as few surface structures are
required.
• Potential for application to all or part of a
given site, depending on site conditions
and cleanup objectives.
• Use in conjunction with, or as a follow-up to
other (active) remedial measures.
Limitations
• There is no single test that can tell if natural
attenuation will occur at a particular site.
• There is a lack of information about the rate
of oxidation and reduction of chromium
under conditions likely to be encountered
by plumes emanating from chromium
sources.
• Longer time frames may be required to
achieve objectives, compared to active
remediation.
• Site characterization may be more complex
and costly.
• Long-term monitoring on a routine basis will
generally be necessary.
• Institutional controls may be necessary to
ensure long-term protectiveness.
• Potential exists for continued contamination
migration, and/or cross-media transfer of
contaminants.
• Hydrologic and geochemical conditions
amenable to natural attenuation are likely
to change over time and could result in
renewed mobility of previously stabilized
contaminants, adversely impacting
remedial effectiveness.
• More extensive education and outreach
efforts may be required in order to gain
public acceptance of monitored natural
attenuation.
3.6.2 Status
Beginning in the late 1980s and continuing
through the 1990s, examination of the
acceptance of natural attenuation at the state
level showed that almost every state was
reviewing its positions with thoughts of
changing them. As of 1996, 38 states were
considering changing their policies toward
acceptance of natural attenuation. The primary
impetus for changing attitudes is the rising costs
for more "active" remediation techniques.
Recently, there has been a slow accumulation
of case histories demonstrating that
groundwater is sometimes better cleaned by
natural processes. Further examination reveals
that in many cases, the health risks do not merit
the expense of engineered remediation (Brady
etal., 1998).
3.6.3 Performance and Cost Data
One of the most successful attempts at
quantifying natural attenuation of Cr(VI) was at
the Trinity Sand Aquifer in Texas. Cr(VI) from a
number of chrome plating operations seeped
into the aquifer between 1969 and 1978. Maps
of the resulting plume (1986 and 1991) were
analyzed to show that nearly three quarters of
the Cr(VI) initially present had been removed,
according to investigators. Fe(ll) and aquifer
organic matter were thought to be the primary
reducing agents. The primary sink for chromium
was thought to be the formation of Cr(OH)3.
Using best estimates for the remaining
hydrologic inputs, it was calculated that
contaminant levels would decrease through
natural attenuation to below MCLs within a
decade (Brady, et al., 1998).
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3.7 Phytoremediation
3.7.1 Technology Description
Phytoremediation uses plants to remediate
contaminated soil and groundwater by taking
advantage of the plants' natural abilities to take
up, accumulate, and/or degrade inorganic and
organic constituents. All plants extract, through
their root systems, necessary nutrients,
including metals from their soil and water
environments. Some plants have the ability to
store large amounts of metals, even some
metals that do not appear to be required for
plant functioning. Metal contaminants that have
been remediated in laboratory and/or field
studies using phytoremediation include Cr(VI),
Cd, Pb, Co, Cu, Ni, Se, and Zn (Miller, 1996).
Phytoremediation technologies are applicable to
sites with low to moderate soil contamination
over large areas, and to sites with large
volumes of groundwater with low levels of
contamination that have to be cleaned to low
standards. They are most effective if soil
contamination is limited to within 3 feet of the
surface, and if groundwater is within 10 feet of
the surface. Groundwater contaminated with
metals can be treated through the use of deep-
rooted trees such as poplars to capture
groundwater, uptake the metals, and retard
contaminant migration (Miller, 1996).
Phytoremediation may be used as a follow-up
technique after areas having high concentration
of pollutants have been mitigated, or in
conjunction with other remediation technologies
(USEPA, 1996).
3.7.2 Status
Phytoremediation technologies are in the early
stage of development, with laboratory research
and limited field trials being conducted to
determine processes and refine methods. Full-
scale remediation projects have not been
completed and regulatory approval may be
difficult to acquire (Miller, 1996). Like
bioremediation and natural attenuation,
mathematical modeling and monitoring are
necessary to demonstrate the effectiveness of
the technology to regulatory agencies (Schnoor,
1997). At the current stage of development, this
process is best suited for sites with widely
dispersed contamination at low concentrations
where only treatment of soils at the surface
(within depth of the root zone) is required. In the
future, phytoremediation may provide a low cost
option under specific circumstances for
treatment of soils contaminated with metals
(U.S. EPA, 1996).
3.7.3 Performance and Cost Data
This technology is attractive because of its
potentially low cost compared to more "active"
remedial approaches. The tradeoff is the
amount of time that is required to achieve
treatment to clean-up levels. Cost estimates for
phytoremediation vary widely. Limited cost and
performance data are currently available. Using
phytoremediation to clean up one acre of sandy
loam to a depth of 50 cm typically will cost
$60,000 to $100,000, compared with a cost of
at least $400,000 for excavation and disposal
storage without treatment (Salt, 1995). The
processing and ultimate disposal of the biomass
generated is likely to be a major percentage of
overall costs, particularly when highly toxic
metals and radionuclides are present at a site
(U.S. EPA, 1996).
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Section 4
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Nyer, E. and P. Palmer. 1997. "In situ
Treatments Using Water as the Carrier."
Environmental Technology-Journal of
Advanced Science and Engineering. Vol 7,
Issue 4. July/August 1997.
Nyer, E. and S. Suthersan. 1996. "Treatment
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Ground Water Monitor Remediation,
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O'Hannesin, S. 1999. "An Overview of
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Presentation for the USEPA Conference
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August, 1999.
Palmer, C. and R. Puls. 1994. Natural
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1994.
Palmer, C.D. and W. Fish. 1992. Chemical
Enhancements to Pump-and-Treat
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Protection Agency Ground Water Issue:
EPA/540/S-92/001. January 1992.
Palmer, C.D. and P.R. Wittbrodt. 1991.
"Processes Affecting the Remediation of
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Environmental Health Perspectives. Vol.
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Pintail Systems, Inc. "Biomineralization of
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2, 1998.
Puls, R. 2000. Review comments on draft
document. July, 2000.
Puls, R.W., C.J. Paul, and R.M. Powell. 1999.
"The Application of In Situ Permeable
Reactive (zero-valent iron) Barrier
Technology for the Remediation of
Chromate-Contaminated Groundwater: a
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Pages 989-1000.
Richard, F.C. and A.C.M. Bourg. 1991.
"Aqueous Geochemistry of Chromium: A
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Pages 807 through 816.
Ross, D.S., R.E. Sjogren, and R.J. Bartlett.
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Toxicity to Microorganisms." Journal of
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148.
Rouse, J.V. 1994. "In-Situ Remediation of
Dissolved Chromate-lon Contamination of
Ground Water." Paper presentation for the
87th Annual Meeting of the Air & Waste
Management Association. June 1994.
Rouse, J.V. 1997. "Natural and Enhanced
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l/l/oocf-Prese/vers' Association. April 27
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Rouse, J.V., and R.Z. Pyrih. 1990. "In-Place
Cleanup of Heavy Metal Contamination of
Soil and Ground Water at Wood
Preservation Sites." Paper Presentation for
the 86th Annual Meeting of the American
Wood-Preservers' Association. April
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Rouse, J.V., M.C. Leahy, and R.A. Brown.
1996. "A Geochemical Way to Keep Metals
at Bay." Environmental Engineering World.
May through June 1996.
Rouse, J.V. Personal communication with Jim
Rouse of Montgomery Watson on August
23, 1999.
Rouse, J., E Benker, M. Daud, and D. Lam.
1999. "In-Situ Remediation of Chromium
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Proceedings, Contaminated Site
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in Freemantle, Western Australia,
sponsored by Centre for Groundwater
Studies, CSIRO Land and Water, March
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Rouse, J.V. 1999a. Monitoring data and results.
Draft Remediation System Progress
Report, South Australia. February 24,
1999.
Sabatini, D.A., R.C. Knox, E.E. Tucker, and
R.W. Puls. 1997. Innovative Measures for
Subsurface Chromium Remediation:
Source Zone, Concentrated Plume, and
Dilute Plume. U.S. Environmental
Protection Agency, Environmental
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September 1997.
Sacre, J. 1997. Treatment Walls: A Status
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Technologies Analysis Center. April 1997.
SAIC (Science Applications International
Corporation). 1998. "Fact Sheet:
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1998.
Salt, D.E., et al. 1995. "Phytoremediation: A
Novel Strategy for the Removal of Toxic
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Biotechnology: Volume 13, May. Pages
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Schmelling, S. 1999. "Overview of Abiotic In
Situ Groundwater Remediation
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Conference on Abiotic In Situ Technologies
for Groundwater Remediation, Dallas,
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Schnoor, J.L. 1997. "Phytoremediation."
Ground-Water Remediation Technologies
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Report. October 1997.
Steimle, R. 1997. "In Situ Flushing Team
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Technologies Analysis Center
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"Adsorption and desorption of hexavalent
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Groundwater Extraction at the United
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Suthersan, S. 1997. Remediation Engineering:
Design Concepts. Lewis Publishers, Boca
Raton. 1997.
Thomasser, R.M. 1999. Pilot study groundwater
monitoring results. Monthly, Quarterly, and
Annual Status Reports to the USEPA,
Region IX, 1999.
Thomasser, R.M. 2000a. Pilot study
groundwater monitoring results. Quarterly
Status Report to the USEPA, Region IX,
July 14, 2000.
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Thomasser, R.M. 2000b. Remedial program
and monitoring results. Combined Fourth
Quarter and Annual Report for 1999.
January 15, 2000.
Thomasser, R.M., and J.V. Rouse. 1999. "In
Situ Remediation of Chromium
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Paper presentation for the American Wood
Preservers Association, May, 1999
Conference on Assessment and
Remediation of Soil and Ground Water
Contamination at Wood Treating Sites.
Tremaine, J. 1999. "In Situ Remediation of
Hexavalent Chromium in Groundwater:
Practical Implementation." Presentation for
the USEPA Conference on Abiotic In Situ
Technologies for Groundwater
Remediation, Dallas, Texas. August, 1999.
U.S. Environmental Protection Agency. 1995. In
Situ Remediation Technology Status
Report: Electrokinetics. EPA 542-K-94-007.
April 1995.
U.S. Environmental Protection Agency. 1995a.
In situ Remediation Technology Status
Report: Treatment Walls. EPA 542-K-94-
004. April 1995.
U.S. Environmental Protection Agency. 1996,
Report: Recent Developments for In Situ
Treatment of Metals-Contaminated Soils.
U.S. Environmental Protection Agency,
Office of Solid Waste and Emergency
Response, draft.
U.S. Environmental Protection Agency. 1996a.
Superfund Innovative Technology
Evaluation Program Technology Profiles,
Ninth Edition. EPA/540/R-97/502.
December 1996.
U.S. Environmental Protection Agency. 1996b.
Permeable Barriers Actions Team. EPA
542-F-96-010C. September 1996.
U.S. Environmental Protection Agency. 1997.
Technology Alternatives for the
Remediation of Soils Contaminated with
As, Cd, Cr, Hg, and Pb. Engineering
Bulletin. EPA/540/S-97/50. August 1997.
U.S. Environmental Protection Agency. 1997a.
Permeable Reactive Subsurface Barriers
for the Interception and Remediation of
Chlorinated Hydrocarbon and Chromium
(VI) Plumes in Ground Water. U.S. EPA
Remedial Technology Fact Sheet.
EPA/600/F-97/008. July 1997.
U.S. Environmental Protection Agency, 1997b.
Electrokinetic Laboratory and Field
Processes Applicable to Radioactive and
Hazardous Mixed Waste in Soil and
Groundwater. EPA/402/R-97/006. USEPA
Office of Radiation and Indoor Air. July
1997.
U.S. Environmental Protection Agency. 1997c.
Treatment Technology Performance and
Cost Data for Remediation of Wood
Preserving Sites. EPA/625/R-97/009.
October 1997.
U.S. Environmental Protection Agency. 1997d.
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Underground Storage Tank Sites."
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U.S. Environmental Protection Agency. 1999a.
An In Situ Permeable Reactive Barrier for
the Treatment of Hexavalent Chromium
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An In Situ Permeable Reactive Barrier for
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the Treatment of Hexavalent Chromium
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1996.
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Department of Energy, Washington. D.C.
PNNL-12153.
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Chemical Treatment. Technology
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the Ground-Water Remediation
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C.C. Ainsworth. 1988. "Chromate
Adsorption by Kaolinite." Clays and Clay
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and C.T. Resch. 1989. "Adsorption of
Chromate by Subsurface Soil Horizons."
So/7 Science Society American Journal.
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Zhaohui, L., H.K. Jones, R.S. Bowman, and R.
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4330.
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APPENDIX A
SOURCES OF ADDITIONAL INFORMATION
This section lists sources that may be useful to the reader to locate additional
information about in situ chromium contamination and remediation technologies. Many
of these sources, but not all of them, were referred to during the preparation of this
guide. Some of these sources may only address organic forms of site contamination.
COMPLETED NORTH AMERICAN INNOVATIVE REMEDIATION TECHNOLOGY DEMONSTRATION
PROJECTS.
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY. EPA 542-B-96-002. AUGUST 12,1996
Analyzes and summarizes information for close to 300 completed demonstration soil and groundwater
remediation projects, including those performed, co-sponsored, or funded through programs
developed by EPA, U.S. military services, DOE, Canadian government, and States of California and
New Jersey. Information includes contaminants treated, technology type, media, vendor, project
sponsor, reports available, and contacts.
ALTERNATIVE TREATMENT TECHNOLOGIES INFORMATION CENTER (ATTIC)
(703)908-2138
WWW.EPA. GOV/ATTIC
Provides access to a collection of hazardous waste databases. Information includes hazardous waste
abstracts, news bulletins, conference information, and a message board. Remediation technology
information is catalogued by contaminant.
CLEANUP INFORMATION (CLU-IN) BULLETIN BOARD
(301) 598-8366
WWW. EPA. GOV/CLUIN
Database with current groundwater remediation technology information. Information includes
bulletins, message and on-file exchange, and on-line databases and directories. Downloadable
documents, abstracts, etc.
VENDOR INFORMATION SYSTEM FOR INNOVATIVE TREATMENT TECHNOLOGIES (VISITT)
(800) 245-4505
WWW. PRCEMI. COM/VISITT
Contains current information about the availability, performance, and cost of innovative technologies
to remediate hazardous waste sites.
WATERLOO HOME PAGE
HTTP://DARCY. UWA TERLOO. CA/HOME. HTML
Only a scaled down version is available now.
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REMEDIATION TECHNOLOGIES DEVELOPMENT FORUM (RTDF)
HTTP://WWW.RTDF. ORG
Contains information on the Bioremediation Consortium, Lasagna Partnership, INERT Soil-Metals
Action Team, Phytoremediation of Organics Action Team, Permeable Reactive Barriers Action Team,
In situ Flushing Action Team, and the Sediments Remediation Action Team.
GROUND-WATER REMEDIATION TECHNOLOGIES ANALYSIS CENTER
HTTP://WWW.GWRTAC. ORG
A national environmental technology transfer center that provides information on the use of
innovative technologies to clean-up contaminated groundwater.
SUPERFUND INNOVATIVE TECHNOLOGY EVALUATION (SITE) PROGRAM
HTTP://WWW.EPA.GOV/ORD/SITE
Contains information on innovative remediation technologies which includes technology profiles,
program highlights, technical documents, and project status.
REMEDIATION AND RESTORATION AT UCLA's CENTER FOR CLEAN TECHNOLOGY
HTTP'J/CCT. SEAS. UCLA.EDU/CCT. RR.HTML
Involves research and development in Transport and Transformation Technology, Simulation and
Management Models, and Risk Assessment and Risk Management Frameworks. Site information
includes brief descriptions of current research.
LAWRENCE LIVERMORE NATIONAL LABORATORY- ENVIRONMENTAL TECHNOLOGIES PROGRAM
HTTP://WWW-EP.ES.LLNLGOV/WWW-EP/AET.HTML
Fundamental and applied research on environmental topics including remediation technologies.
Remediation focus areas include (1) in situ thermal remediation, including Dynamic Underground
Stripping and in situ electrical heating methods, and (2) in situ bioremediation, including an in situ
Microbial Filter method. Information provided about these technologies include descriptions of
concepts and processes, field test results, and points of contact. Science and technology articles,
including remediation issues, available on-line.
PACIFIC NORTHWEST NATIONAL LABORATORY (PNNL) TECHNOLOGY ABSTRACTS
HTTP://W3. PNL G O V:2080/TRANSFER/T2HOME. HTML
On-line technology briefs and fliers about technologies available for licensing from PNNL.
Technologies include in situ bioremediation, in situ vitrification. Remedial Action Assessment
Software (RAAS), and in situ corona.
PACIFIC NORTHWEST LABORATORIES PRO TECH (PROSPECTIVE TECHNOLOGIES) ONLINE
HTTP://GII-AWARDS.COM/NICAMPGN/244A.HTML
ProTech is a graphical communication application used by DOE's Office of Technology Development
to describe information on over 150 innovative environmental technologies being developed and
demonstrated at DOE sites. Information provided includes technical summaries of ongoing or planned
site characterization and remediation technology applications at DOE facilities.
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USEPA ROBERT S. KERR ENVIRONMENTAL RESEARCH LABORATORY, NOW KNOWN AS SUBSURFACE
PROTECTION REMEDIATION DIVISION (SPRD)
HTTP://WWW. EPA. G O V/ADA/KERRLAB. HTML
Bibliography with abstracts of scientific and technical publications developed by and through the
SPRD relating to groundwater protection, fate and transport, and remediation. Information packets on
bioremediation and other related topics are also available. Houses the Center for Subsurface
Modeling Support (CSMoS), which provides free groundwater modeling software and services.
USDOE OFFICE OF ENVIRONMENTAL MANAGEMENT
HTTP://WWW. EM. DOE.GOV
Information about DOE's technology needs in the areas of characterization, storage, and disposal of
hazardous radioactive wastes. Information on technology at DOE sites. Directory of technologies and
points of contacts relating to sites and technologies.
CANADIAN CENTER FOR INLAND WATERS GROUNDWATER REMEDIATION PROJECT
HTTP://GWRP. caw. CA/GWRP
Bibliography and abstracts of recent publications, covering groundwater remediation and related
topics. Groundwater modeling and analysis software is also available.
GLOBAL NETWORK FOR ENVIRONMENTAL TECHNOLOGIES
HTTP://WWW. GNET. ORG
Contains extensive information under the "Technology Center" subsite.
NATIONAL HYDROLOGY RESEARCH INSTITUTE (NHRI) ENVIRONMENT CANADA
HTTP://GWRP.CCIW.CA/NHRI/NHRIGW.HTML
Investigation of natural processes in subsurface environments and development of remedial
techniques for cleanup and containment of toxic contaminants. Research briefs include topics such as
the effects of microbial biofilms on organic contaminants and in situ barriers for groundwater
remediation and containment.
GROUNDWATER AND SITE REMEDIATION CSIRO-AUSTRALIA, INCLUDING CENTER FOR
GROUNDWATER STUDIES
HTTP://WWW.DWR. CSIRO.AU
Brief description of current research and a bibliographic listing of publications (1985-1996) on topics
such as natural BTEX biodegredation, enhanced bioremediation, trichloroethylene, munitions, aquifer
bioclogging, and others.
HAZARDOUS SUBSTANCE RESEARCH CENTER
HTTP://WWW. G TRI. GA TECH. EDU/HSRC
Research projects involve finding innovative technologies to remediate hazardous organic
contaminants. Research topics include in situ bioremediation, surfactants, and bioventing. Also
contains an accessible site and infrastructure for filed projects. Includes research project abstracts,
research briefs, points of contact, and other publications.
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UNIVERSITY OF IDAHO CENTER FOR HAZARDOUS WASTE REMEDIATION RESEARCH
HTTP://UIDAHO. EDU/-CRA WFORD/CENTER2. HTML
Focuses on the characterization of contaminated sites and the development and field application of
novel technologies for hazardous waste remediation. Remediation technologies addressed include
bioremediation and geochemical remediation. Site includes a bibliography of representatives.
STANFORD HYDROGEOLOGY RESEARCH GROUP
HTTP://PANGEA.STANFORD.EDU/HYDRO/HYDRO.HTML
Current research topics include in-Well Vapor Stripping, Optimal Remedial Design, and Aquifer
Heterogenicity. Information about these topics includes research abstracts and references.
WATERNET
HTTP://WATERNET. COM
Articles from Water Technology and International Ground Water Technology journals, including
remediation and related topics.
CENTER FOR GROUNDWATER RESEARCH (CGR) OREGON GRADUATE INSTITUTE OF SCIENCE AND
TECHNOLOGY (OGI)
HTTP://WWW.ESE.OGI_DOCS/CGR.HTML
Provides brief descriptions and points of contact for research activities being conducted at OGI,
including "Research on Contaminant Remediation with Zero-Valent Iron Metal."
NATIONAL INSTITUTE OF ENVIRONMENTAL HEALTH SCIENCES SUPERFUND BASIC RESEARCH
PROGRAM (SBRP)
HTTP://NIEHS. NIH. G o V/SBRP/HOME. HTML
Presents brief highlights of their research activities, at universities across the country, in areas
including remediation and bioremediation.
BERKELEY ENVIRONMENTAL RESTORATION CENTER
HTTP://BERC3. ME. BERKELEY. EDU
Includes abstracts of student research in the areas of remediation and fate and transport.
THE CENTER FOR ENVIRONMENTAL BIOTECHNOLOGY (CEB)
HTTP://WWW. RA. UTK. EDU
Information includes a list and brief descriptions of current research projects, and contacts for further
information.
USEPA OFFICE OF RESEARCH AND DEVELOPMENT
Bulletin Board Service (513) 569-7610
Provides a bibliography of over 19,000 documents and a message board.
CENTER FOR ENVIRONMENTAL RESEARCH INFORMATION (CERI) - TECHNOLOGY TRANSFER
WWW. EPA. GOV/TTBNRMRL
Provides information and documents concerning the latest developments for environmental
technologies and problems including monitoring, treatment, and other research.
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DIALOG DATABASE
(800) 3-DIALOG
Contains files relevant to hazardous waste including Enviroline, CA Search, Pollution Astracts,
Compendex, Energy Science and Technology, National Technical Information Service (NTIS), etc.
NTIS Database contains abstracts of government-sponsored research, development, and engineering
analysis prepared by approximately 250 Federal agencies and some state and local governments.
SOIL AND GROUNDWATER MAGAZINE (FORMERLYSOILS MAGAZINE)
HTTP://WWW. GVI.NET
Current and back issues containing relatively brief articles on remediation and related topics.
NATURAL ATTENUATION: CERCLA, RBCA 's, AND THE FUTURE OF ENVIRONMENTAL REMEDIATION
PATRICK V. BRADY, MICHAEL V., BRADY, DAVID J. BORNS. LEWIS PUBLISHERS, 1998
Book analyzes the historical evolution and current direction of environmental remediation in the
United States and outlines why there is now and will be an increasing reliance on natural attenuation
of hazardous substance toxicity in the coming years.
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APPENDIX B
TECHNOLOGY AND VENDOR CONTACT INFORMATION
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TECHNOLOGY AND VENDOR CONTACT INFORMATION
Technology/Vendor
Contact
Person
Address
Geochemical
Fixation/Montgomery
Watson
Jim Rouse
370 Interlocken Blvd.
Suite 300
Broomfield, CO 80021
303-410-4029
Jim.V.Rouse@mw.com
Geochemical Fixation
Ralph Howard,
RPM
EPA Region IV
Atlanta Federal Center
100 Alabama Street, S.W.
Atlanta, Georgia 30303
404-562-8829
howard.ralph@epa.gov
PRBs
Dr. Robert Puls
EPA
Subsurface Protection and
Remediation Division
National Risk Management
Research Laboratory
Ada, OK 74820
580-436-8543
puls.robert@epa.gov
PRBs/Waterloo Centre
forGroundwater
Research
Dr. David Blowes
or Dr. Robert
Gillham
Department of Earth
Sciences
University of Waterloo
Ontario, N2L 3G1, Canada
519-888-4878
blowes@sciborg.uwaterlo
o.ca
519-888-4658
rwgillha@sciborg.uwaterlo
o.ca
PRBs/ISRM
Dr. John
Fruchter
Battelle Pacific NW Lab
PO Box 999
Richland, WA 99352
509-376-3937
iohn.fruchter@pnl.gov
PRBs/Zeolite & ZVI
Dr. Robert
Bowman
Dept. Of Earth and
Environmental Science
New Mexico Tech
801 Leroy Place
Socorro, NM 87801
505-835-5992
bowman@nmt.edu
PRBs/Cercona of
America, Inc.
Richard Helferich
President
5911 Wolf Creek Pike
Dayton, OH 45426
937-854-9860
rhelferich@coax.net
Reactive
Zones/Arcadis
Geraghty & Miller
Dr. Suthan
Sutherson or
Frank Lenzo
3000 Cabot Blvd. West
Suite 300
Langhorne, PA 19047
215-752-6840
ssuthers@gmgw.com
flenzo@gmgw.com
Biomineralization/
Pintail Systems, Inc.
Leslie Thompson
11801 E. 33rd Avenue
Suite C
Aurora, CO 80010
303-367-8443
Soil Flushing &
Chromium Extraction
Dr. David
Sabatini
Institute for Applied
Surfactant Research
University of Oklahoma
Norman, OK 73019
405-325-4273
sabatini@mailhost.ecn.uo
knor.edu
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Technology/Vendor
Soil Flushing &
Chromium Extraction
Electrokinetics
Electro kinetics/Geoki-
netics International,
Inc.
Electokinetics/
Electrokinetics, Inc.
Electro kinetics/Sandia
National Laboratories
Natural Attenuation
Phytoremediation/
Phytotech, Inc.
Contact Person
Richard Steimle
Randy Parker
Robert Clarke or
Stuart Smedley
Elif Acar or
Robert Gale
Eric Lingren
Dr. Robert Puls
Michael Blaylock
or John Ehrler
Address
U.S. EPA
Technology Innovation
Office
5702G, 401 M Street SW
Washington DC 20460
U.S. EPA
National Risk Management
Research Laboratory
26 W. Martin Luther King
Dr.
Cincinnati, OH 45268
829 Heinz Street
Berkeley, CA 94710
11552 Cedar Park Ave.
Baton Rouge, LA 70809
P.O. Box5800
Albuquerque, NM 87185
U.S. EPA
Subsurface Protection and
Remediation Division
National Risk Management
Research Laboratory
Ada, OK 74820
One Deer Park Drive
Suite 1
Monmouth Junction, NJ
08852
703-603-7195
steimle. rich(S>,epa.qov
513-569-7271
parker.randy@epa.qov
510-704-2940
edarobert@aol.com
504-753-8004
ekinc@pipeline.com
505-844-3820
erlindq@sandia.gov
580-436-8543
puls.robert@epa.gov
908-438-0900
soilrx@aol.com or
iohnehrler@aol.com
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