INTRODUCTION TO IN SITU
BIOREMEDIATION
OF GROUNDWATER
Groundwater
to Recirculation
System
Groundwater
Flow Direction
Bedrock
&EPA
United States
Environmental Protection
Agency
Office of Solid Waste and Emergency Response
EPA542-R-13-018
December 2013
-------
&EPA
542-R-13-018
United States December 2013
Environmental Protection Office of Solid Waste and Emergency Response
Agency
INTRODUCTION TO IN SITU
BIOREMEDIATION OF GROUNDWATER
EXECUTIVE SUMMARY
Bioremediation is an engineered technology that modifies environmental conditions (physical, chemical,
biochemical, or microbiological) to encourage microorganisms to destroy or detoxify organic and
inorganic contaminants in the environment. The process can be applied above ground in land farms,
tanks, biopiles, or other treatment systems (referred to as ex situ) or below ground in the soil or
groundwater, referred to as in situ. In situ bioremediation of groundwater has become one of the most
widely used technologies for contaminated site treatment because of its relatively low cost, adaptability
to site-specific conditions, and efficacy when properly implemented (Stroo 2010).
Introduction to In Situ Bioremediation of Groundwater was prepared by the Office of Superfund
Remediation and Technology Innovation (OSRTI) as an introduction to in situ bioremediation of
groundwater. This information is intended for U.S. Environmental Protection Agency (EPA) and state
agency site managers and may serve as a reference to designers and practitioners. Others may find the
EPA's Citizen's Guide to Bioremediation (EPA 2012a) to be a more fundamental and concise reference.
In situ bioremediation (ISB) of groundwater involves the encouragement of indigenous bacterial
populations to metabolize target contaminants through the addition of various amendments
(biostimulation) to the subsurface environment. In addition to amendments, select strains of bacteria
may be added to the subsurface to help treat some sites (bioaugmentation). Bacteria perform coupled
oxidation/reduction (redox) reactions to live, and bioremediation exploits these reactions to remove
contaminants from contaminated media (soil, air, or groundwater). Bacteria can use different electron
acceptors (oxidized compounds) and donors (reduced compounds) in the three major oxidation
pathways — aerobic respiration, anaerobic respiration, and fermentation. ISB can use all of these
pathways, and contaminant degradation may occur through direct metabolism, cometabolism, or
abiotic transformations that may result from biological activities.
Aerobic bioremediation most commonly takes place in the presence of oxygen and relies on the direct
microbial metabolic oxidation of a contaminant. The primary concern when an aerobic bioremediation
system is designed is delivery of oxygen, which is the electron acceptor. Aerobic bioremediation is most
effective in treating non-halogenated organic compounds. Many reduced contaminants can be
aerobically degraded by aerobic bacteria already present in the subsurface environment. Oxygen can be
added directly to the subsurface, or chemical oxidants can be applied, which release oxygen as they
dissolve or decompose. Oxygen and oxygen-releasing compounds can be delivered to the groundwater
via several methods, depending on their physical properties, site hydrogeology, and the desired delivery
efficiency. The end products of aerobic respiration are usually carbon dioxide and water.
Introduction to In Situ Bioremediation of Groundwater ES-1
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Anaerobic oxidative bioremediation takes place in the absence of oxygen. It relies on other electron
acceptors such as nitrate or sulfate for direct microbial metabolic oxidation of a contaminant. This
approach is often applied at petroleum-contaminated sites where oxygen has already been depleted.
Amendments with soluble sulfate and electron donor are often added to the affected area to stimulate
sulfate-reducing conditions to help microbes metabolize the petroleum compounds. A byproduct of this
approach is hydrogen sulfide. The hydrogen sulfide can react with the iron at sites where metals such as
iron occur naturally to produce iron sulfide or pyrite and reduce the amount of hydrogen sulfide.
Anaerobic reductive bioremediation takes place in the absence of oxygen. It relies on the presence of
biologically available organic carbon, which may be naturally present or added to stimulate activity. The
organic carbon, also commonly called an organic substrate or an electron donor source, creates and
sustains anaerobic conditions by consuming oxygen and other electron acceptors during its
biodegradation. It also promotes the bioreduction of oxidized contaminants such as chlorinated solvents
(EPA 2001b) by generating hydrogen through fermentation reactions. Because these contaminants exist
in an oxidized state, they are generally much less susceptible to aerobic oxidation processes, but they
can be reduced by microbes under anaerobic conditions, a process also referred to as enhanced
reductive dechlorination (ERD) when applied to chlorinated solvents. In many cases, microorganisms use
the oxidized contaminants in a respiratory mechanism and are able to derive metabolically useful
energy (EPA 2000, AFCEE 2004). Anaerobic conditions may be used to degrade highly chlorinated
contaminants, such as tetrachloroethene (PCE) and trichloroethene (TCE) to ethene, 1,1,1-
trichloroethane (1,1,1-TCA) to ethane, carbon tetrachloride (CT) to methane, or perchlorate to chloride
and oxygen. Microbially induced reduction of hexavalent chromium to trivalent chromium may be the
most common application of bioremediation to metals.
Cometabolism occurs when microorganisms using one compound as an energy source fortuitously
produce an enzyme that chemically transforms another compound. Organisms thus can degrade a
contaminant without gaining any energy from the reaction. Cometabolic degradation is a process that
often happens concurrently in bioremediation systems designed for direct metabolism of contaminants;
however, some systems have been designed to specifically take advantage of cometabolic processes.
Hazen (2009) indicates that cometabolic bioremediation can occur in environments where contaminant
concentrations are well below concentrations that could provide a carbon or energy benefit to the
biodegrader. Therefore, this method may be effective at degrading very low concentrations of some
contaminants.
Adequate site characterization is critical to designing a successful ISB remedy. The nature and extent of
the environmental impacts need to be known, as well as several key characteristics of the affected
media. Development of a conceptual site model (CSM) helps guide the characterization and subsequent
design, implementation, and operation of the remedy. Application of ISB is highly dependent on site
characteristics, such as the aquifer type, baseline geochemistry, and lithology. Bioremediation can
change site geochemistry by altering the pH or redox status and, as a result, produce secondary
contaminants. Production of secondary contaminants is fairly well understood and is often addressed as
part of the design.
Introduction to In Situ Bioremediation of Groundwater ES-2
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The design process commonly includes bench- and pilot-scale treatability studies. These tests may be
performed during the feasibility study or remedial design, are used to evaluate whether the proposed
bioremediation remedy will be successful, and obtain important design criteria. In general, full-scale
implementation is based on the site CSM, remedial objectives, regulatory requirements, and future site
use or development. The three primary approaches to full-scale implementation for ISB are classified as
active, semi-passive, and passive treatment, distinguished by the need for active groundwater
recirculation during operations (Stroo and Ward 2009). Once designed and installed, a bioremediation
system requires careful monitoring and possible modifications to optimize performance.
Implementation costs related to almost any technology increase with greater depth and greater
treatment volume. Reapplication of amendments, including electron acceptors or donors, will be
required at most sites. Some sites may require geochemical adjustment and nutrient amendment. The
success of biological technologies is highly dependent on the delivery and longevity of the amendments
added to the site and requires a comprehensive performance monitoring program.
This document also highlights several recent trends affecting ISB. These trends include the increasing
emphasis on green or sustainable remediation, the use of stable isotopes as diagnostic tools, high-
resolution site characterization (HRSC), and three-dimensional visualization and analysis (3DVA) of site
data. ISB often results in a smaller on-site environmental footprint than ex situ or non-biological
methods because of its relatively low energy use and the minimal equipment and site disruption
required to implement it.
Stable isotope analysis can be used to demonstrate that biodegradation is occurring, to discriminate
between biological and nonbiological processes and to estimate the rate and extent of contaminant
degradation. Stable isotope probing, where compounds enriched in a stable isotope are added to the
subsurface, is being used to measure the fraction of degradation directly caused by microbial activity
during bioremediation.
Several molecular biological tools (MBTs) are becoming more widely available and cost effective for
applications in support of site characterization, remediation, monitoring, and closure, which include but
are not limited to microassays, Fluorescence In Situ Hybridization (FISH), and quantitative Polymerase
Chain Reaction (qPCR) (ITRC 2013). MBTs can be used to determine if the necessary bacteria with the
right genes are present at the site and if they exist at optimal levels.
All remedial technologies, including ISB, require accurate site characterization techniques, such as HRSC.
HRSC has become more prominent as sampling and analytical techniques, data evaluation, and
visualization methods have improved. HRSC strategies and techniques use scale-appropriate
measurement and sample density to define contaminant distributions, and the physical context in which
they reside, with greater certainty, supporting faster and more effective site cleanup (CLU-IN 2013).
Lithologic, hydrogeologic and contaminant data are often provided by real-time direct sensing and
hydraulic profiling technologies such as Laser Induced Fluorescence (LIF), Membrane Interface Probes
(MIP), and electrical conductivity (EC) probes. Several software programs are now available to perform
3DVA of site characterization and performance monitoring data. These programs are useful for
Introduction to In Situ Bioremediation of Groundwater ES-3
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designing amendment delivery systems and determining which portions of a plume may require
additional amendments. The programs typically use geostatistical kriging procedures to establish the
spatial distribution of each parameter in three-dimensional space.
Notice and Disclaimer:
Preparation of this report has been funded by the U.S. Environmental Protection Agency Office of
Superfund Remediation and Technology Innovation (OSRTI) under Contract Number EP-W-07-078 to
Tetra Tech EM Inc. This report is not intended, nor can it be relied on, to create any rights enforceable
by any party in litigation with the United States. Mention of trade names or commercial products does
not constitute endorsement or recommendation for use. A PDF version of Introduction to In situ
Bioremediation ofGroundwater is available for viewing or downloading from the Hazardous Waste
Cleanup Information (CLU-IN) system website at www.clu-in.org. For questions concerning this
document, contact Linda Fiedler (703-603-7194/fiedler.linda@epa.gov) or Edward Gilbert (703-603-
8883/gilbert.edward@epa.gov ), Office of Superfund Remediation and Technology Innovation, U.S.
Environmental Protection Agency.
Introduction to In Situ Bioremediation ofGroundwater ES-4
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Table of Contents
Executive Summary 1
Acronyms and Abbreviations iv
1.0 Introduction 1
1.1 Superfund Project Information 2
1.2 History and Background 2
1.3 Microbiology 5
1.4 Reduction and Oxidation Chemistry and Microbial Metabolism 5
1.4.1 Aerobic Respiration 6
1.4.2 Anaerobic Respiration and Fermentation 6
1.4.3 Direct Metabolism 7
1.4.4 Cometabolism 8
1.4.5 Abiotic Transformation 8
1.5 Conceptual Site Model 9
1.5.1 Land Use and Risk 10
1.5.2 Geologic Setting 11
1.5.3 Hydraulic Properties of Contaminated Media 11
1.5.4 Biogeochemistry 13
1.5.5 Contaminant Distribution 19
2.0 Strategies for Groundwater Bioremediation 23
2.1 Aerobic Bioremediation 23
2.1.1 Common Applicable Contaminants 23
2.1.2 Key Microbe Summary 23
2.1.3 Sources of Electron Acceptor 24
2.1.4 Delivery Mechanisms 24
2.1.5 Common Byproducts 26
2.2 Anaerobic Oxidative Bioremediation 26
2.2.1 Common Applicable Contaminants 27
2.2.2 Key Microbe Summary 27
2.2.3 Sources of Electron Acceptor 27
2.2.4 Delivery Mechanisms 28
Introduction to In Situ Bioremediation of Groundwater i
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Table of Contents, (continued)
2.2.5 Common Byproducts 28
2.3 Anaerobic Reductive Bioremediation 29
2.3.1 Common Applicable Contaminants 30
2.3.2 Key Microbe Summary 32
2.3.3 Sources of Electron Donor 33
2.3.4 Delivery Mechanisms 35
2.3.5 Common Byproducts 36
2.4 Cometabolic Bioremediation 36
3.0 Field Implementation 38
3.1 Treatability Studies 38
3.1.1 Bench Test 38
3.1.2 Pilot Test 39
3.2 Full-scale Implementation 41
3.2.1 Active Treatment Approach 42
3.2.2 Semi-Passive Treatment Approach 43
3.2.3 Passive Treatment Approach 43
3.2.4 Vertical Application and Distribution 45
3.2.5 Maintenance 46
3.3 Measuring Performance 47
4.0 Emerging Trends 50
4.1 Environmental Remedy Footprint 50
4.2 Compound Specific Isotope Analysis 50
4.3 High-Resolution Site Characterization 51
4.4 3-D Visualization and Analysis of Data and In Situ Sensors 51
5.0 Summary 53
6.0 References... ...55
Introduction to In Situ Bioremediation of Groundwater
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Figures
Figure 1. Use of In Situ Groundwater Bioremediation Technologies at Superfund Sites 3
Figure 2. Aerobic and Anaerobic Bioremediation Projects atNPL Sites 4
Figure 3. Contaminant Groups Addressed by Bioremediation Technologies at Superfund Sites 4
Figure 4. Dominant Terminal Electron-Accepting Process, Electron Acceptors, and Typical Chemical
Species Responses (Modified from AFCEE 2004, Bouwer and McCarty 1984) 7
Figure 5. Example Biotic and Abiotic Degradation Pathways of Common CVOCs (EPA 2009,
O'Loughlin and Burris 2004) 9
Figure 6. Conceptual Site Model - Exposure Pathway Schematic (ATSDR2005) 10
Figure 7. Conceptual Matrix Diffusion Mechanism (Adapted fromNRC 2004) 13
FigureS. Estimated ORP of Commonly Monitored Species (ITRC 2005) 15
Figure 9. Average Temperature of Shallow Groundwater across the continental U.S. (EPA 2013,
Ecosystems Research, Athens, GA) 17
Figure 10. Permeable Reactive Barrier Example (EPA 200Ib) 35
Figure 11. In situ Bioremediation System Configurations (EPA 2000), as adapted 40
Figure 12. Bioremediation Design Types 41
Figure 13. Typical Circulation System Layout (Kovacich and others 2006) 42
Figure 14. Schematic of Source Area and Barrier Injection Configurations. Adapted from AFCEE 200443
Figure 15. Example Cost Comparison for a PRB with Various Injection Well Spacings (ESTCP 2006).. 44
Figure 16. Example Injection Well Design. (Courtesy of TetraTech, Inc.) 46
Figure 17. Example of Geostatistical Kriging Analysis of Multi-depth TCE Soil Concentrations.
(Courtesy of TetraTech, Inc.) 52
Tables
Table 1. Substrates used for enhanced anaerobic bioremediation
(modified from ITRC 2008, AFCEE 2004) 33
Table 2. Common cometabolic bioremediation substrates, enzymes, and contaminants
(from Hazen 2009) 37
Table 3. Summary of ISB strategies 54
Appendix
Appendix A: A Selection of Superfund Program In Situ Groundwater Bioremediation Sites (Remedies
Selected FY 1989 to 2008)
Introduction to In Situ Bioremediation of Groundwater
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ACRONYMS AND ABBREVIATIONS
3DVA
AFCEE
AST
BTEX
CAM
CERCLA
CLU-IN
COC
CSIA
CSM
CT
CVOC
DCA
DNA
DCE
DNAPL
DPRB
EPA
ERD
ESD
ESTCP
EVO
FISH
gpm
HRSC
ISB
ITRC
LIF
LNAPL
MBT
mg/L
MIP
MTBE
mV
MVS
NAPL
NDMA
NPL
NRC
three-dimensional visualization and analysis
Air Force Center for Engineering and the Environment
Aboveground Storage Tank
Benzene, toluene, ethyl benzene, xylene
Chlorinated aliphatic hydrocarbons
Comprehensive Environmental Response, Compensation, and Liability Act
Cleanup Information
Contaminant of concern
Compound specific isotope analysis
Conceptual site model
Carbon tetrachloride
Chlorinated volatile organic compound
Dichloroethane: 1,1- and 1,2- isomers
Deoxyribonucleic acid
Dichloroethene: cis-1,2-; trans-1,2-; and 1,1- isomers
Dense non-aqueous phase liquid
Dissimilatory perchlorate-reducing bacteria
U.S. Environmental Protection Agency
Enhanced reductive dechlorination
Explanation of Significant Differences
Environmental Security Technology Certification Program
Emulsified vegetable oil
Fluorescence In Situ Hybridization
Gallons per minute
High-resolution site characterization
In situ bioremediation
Interstate Technology and Regulatory Council
Laser Induced Fluorescence
Light non-aqueous phase liquid
Molecular biological tools
Milligrams per liter
Membrane Interface Probe
Methyl tert-butyl ether
Millivolts
Mine Visualization System
Non-aqueous phase liquid
A/-Nitrosodimethylamine
National Priorities List
Nuclear Regulatory Commission
Introduction to In Situ Bioremediation of Groundwater
IV
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O&M
ORP
OSTRTI
OSWER
PAH
PCB
PCE
PCP
ppm
PRB
psi
qPCR
RDX
Redox
ROD
RPM
SIP
SVOC
TCA
TCE
TNT
UST
VC
vcrA
VOC
Operation & Maintenance
Oxidation-reduction potential
Office of Superfund Remediation and Technology Innovation
Office of Solid Waste and Emergency Response
Polycyclic aromatic hydrocarbons
Polychlorinated biphenyls
Perchloroethene or tetrachloroethene
Pentachlorophenol
Parts per million
Permeable reactive barriers
Pounds per square inch
Quantitative Polymerase Chain Reaction
Royal Demolition Explosive
Oxidation/reduction
Record of Decision
Remedial project managers
Stable isotope probing
Semi volatile organic compound
1,1,1-Trichloroethane
Trichloroethene
Trinitrotoluene
Underground Storage Tank
Vinyl chloride
Vinyl chloride reductase
Volatile organic compound
Introduction to In Situ Bioremediation of Groundwater
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1.0
Introduction to In Situ Bioremediation ofGroundwater was prepared by the Office of Superfund
Remediation and Technology Innovation (OSRTI) as an introduction to in situ bioremediation (ISB) of
groundwater. This information is intended for U.S. Environmental Protection Agency (EPA) and state
agency site managers and may serve as a reference to designers and practitioners. Others may find the
EPA's Citizen's Guide to Bioremediation (EPA 2012a) to be a more fundamental and concise reference.
Bioremediation is an engineered technology that modifies environmental conditions (physical, chemical,
biochemical, or microbiological) to encourage microorganisms to detoxify organic and inorganic
contaminants in the environment. The process can be applied above ground in land farms, stirred tanks,
biopiles, or other units (referred to as ex situ) or below ground in the soil or groundwater, referred to as
in situ ("in place") treatment.
This document focuses specifically on in situ groundwater bioremediation. In the context of this
document, groundwater remediation is defined as remediation of contaminants that exist below the
water table. As a result of phase equilibrium in the subsurface, groundwater remediation must address
contaminants dissolved in groundwater as well as those sorbed to the aquifer matrix to be effective. In
some cases, even treatment of non-aqueous phase liquid (NAPL) may be needed. Consideration must
also be given to the capillary fringe and the smear zone, which can serve as an ongoing source of
contaminants to groundwater. This report does not discuss phytoremediation (use of plants to treat
groundwater and soil) or monitored natural attenuation (a technology based on monitoring the progress
of natural, non-engineered processes that often include biodegradation). Those readers interested in
more information on ex situ bioremediation, bioremediation of soil, monitored natural attenuation or
phytoremediation may find useful information on EPA's CLU-IN website (www.cluin.org).
The document provides technical information on evaluating and implementing in situ groundwater
bioremediation at contaminated sites. Superfund program guidance for selecting and implementing
groundwater remedies at Superfund sites can be found in numerous program guidance documents, such
as:
• Office of Solid Waste and Emergency Response (OSWER) 9283.1-33, Summary of Key Existing
Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) Policies for
Groundwater Restoration, June 2009; and
• OSWER 9283.1-34, Groundwater Road Map: Recommended Process for Restoring Contaminated
Groundwater at Superfund Sites, July 2011.
For other groundwater guidance, refer to the Superfund groundwater website
(www.epa.gov/superfund/health/conmedia/gwdocs).
This document provides (1) a brief technical background on ISB, (2) a summary of the use of ISB for
various contaminants, including information on its use for Superfund sites, (3) considerations for
implementation of ISB, (4) brief summaries of some important emerging trends affecting ISB, and
(5) links to additional sources of information.
Introduction to In Situ Bioremediation ofGroundwater
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1.1 Superfund Project Information
The list of in situ groundwater bioremediation projects accompanying this report in the appendix was
derived primarily from the lists in Treatment Technologies for Site Cleanup: Annual Status Report,
Twelfth Edition, and its successor Superfund Remedy Report, Thirteenth Edition, and represents a subset
of in situ groundwater bioremediation projects at National Priorities List (NPL) sites. Projects on the list
include remedial actions for in situ groundwater bioremediation selected in Superfund Records of
Decision (RODs), ROD amendments, and Explanations of Significant Differences (ESDs) for fiscal years
1989 through 2008. These documents are referred as "decision documents." Although decision
documents select a general technology such as ISB, the final selection of a specific design is typically
deferred to the remedial design phase. Detailed information regarding the remedial design and
contaminants treated was compiled for each project on the list based on documents available either on
line (for instance, http^ or in site files. These
sources included 5-year reviews, Superfund site summary fact sheets, remedial action reports, and other
pertinent documents. Remedial project managers (RPMs) and contractors were contacted for additional
clarification as needed. Status information for most projects was last updated in November 2011. More
information on project implementation status, design, and performance may be available on the
websites related to each site found at the link given above.
1.2 History and Background
Bioremediation is not a new concept. Biological treatment of domestic wastewater has been in use since
the mid-1800s, and land treatment has been used for several decades to treat oil and other petroleum
wastes by aerobic biodegradation (Loehr 1979). The basic principles and experience from these
technologies were adapted to ISB of petroleum (and other contaminants) in the 1980s (Thomas and
Ward 1989). ISB has been further developed to treat a wide variety of other contaminants, particularly
since the early 1990s, when the potential for enhanced anaerobic treatment became clear (NRC 1993;
Alexander 1994).
The first use of ISB was in 1972, when aerobic treatment was used to clean up a Sun Oil pipeline spill in
Ambler, Pennsylvania. Treatment consisted of withdrawing groundwater, adding oxygen and nutrients,
and recirculating it through the subsurface (Raymond 1977). Aerobic biological treatment or oxidation of
petroleum releases gained acceptance throughout the 1970s and 1980s and has been used in several
large-scale applications, including the effort to clean up numerous Superfund sites (see for example EPA
1989; Ross 1988).
Anaerobic bioremediation gained popularity when it was recognized as an effective method to
remediate chlorinated solvents in groundwater. In 1997 scientists isolated a bacterium originally
referred to as Dehalococcoides ethenogenes strain 195, the first organism known to completely
dechlorinate the common groundwater contaminant perchloroethene (PCE, also known as
tetrachloroethene) (Maymo-Gatell 1997). Further studies showed that several related bacteria, all now
referred to as strains of Dehalococcoides mccartyi (Loffler and others 2012), had the ability to partially
or completely dechlorinate PCE and the related chloroethenes. To date, these are the only known
organisms with the ability to completely degrade these compounds, which are particularly prevalent
groundwater contaminants at Superfund sites. As a result, several demonstration-scale applications of
Introduction to In Situ Bioremediation of Groundwater 2
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anaerobic bioremediation were completed in the late 1990s and early 2000s. Several of the
demonstration projects went full scale, and today reductive dechlorination, as it is now known, is a
widely accepted method for treating halogenated ethenes, ethanes, and methanes (Stroo 2010).
Figure 1 shows the surge in popularity of anaerobic bioremediation for use at NPL sites after the method
was successfully demonstrated in the early 2000s, while the use of aerobic bioremediation has remained
relatively steady. As indicated in Figures 1 and 2, the selection of anaerobic bioremediation to remediate
groundwater at Superfund sites increased dramatically over recent years, and this method is now used
at the majority of Superfund sites where ISB technologies have been selected.
As shown in Figure 3, the most common groundwater contaminants addressed by ISB at NPL sites were
halogenated volatile organic compounds (VOCs); followed by nonhalogenated VOCs; nonhalogenated
semi-volatile organic compounds (SVOCs); and BTEX compounds (benzene, toluene, ethyl benzene and
xylenes).
Unknown Type
i Anaerobic
i Aerobic
CTlffiffiO^oSCTioSoSeTioSffiOOQ
Fiscal Year Selected
Note: Refer to Appendix for data
Figure 1. Use of In Situ Groundwater Bioremediation Technologies at Superfund Sites.
Introduction to In Situ Bioremediation of Groundwater
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In Situ Anaerobic
Reduction
81
70%
Aerobic
31
Biosparging 27%
\
Note: Includes remedies selected in FY 1989-2008
decision documents. Refer to Appendix for data.
Unknown Type
4
3%
Figure 2. Aerobic and Anaerobic Bioremediation Projects at NPL Sites.
I 1 1 I H
Unknown Type
I Anaerobic
I Aerobic
/ 3?
Note: Includes remedies selected in
FY 1989-2008 decision documents.
Refer to Appendix for data
Figure 3. Contaminant Groups Addressed by Bioremediation Technologies at Superfund Sites.
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1.3 Microbiology
One component of designing an effective ISB system is to understand the fundamental ecology and
physiology of microbes. Microbes have been found everywhere on earth, including environments of
extreme heat, cold, and pH, without oxygen, and in the presence of radiation. They are adaptive,
resilient, and can thrive in environments impaired by most contaminants. Bioremediation most
commonly uses bacteria for treatment, but also includes remediation performed by archaea, protists,
and fungi. Microbes used for bioremediation are often referred to collectively as "bacteria," or "bugs" in
the bioremediation field.
All microbes have basic requirements for life and growth. The six elements considered essential for life
are carbon, hydrogen, oxygen, nitrogen, phosphorus, and sulfur. The needs of bacteria can be further
simplified to three requirements: a carbon source that can be used to build its biomass, an electron
donor (such as hydrogen) for the energy it needs to live and reproduce, and a terminal electron acceptor
(for example, oxygen) to receive the electrons the bacteria use for energy. Often, the carbon source will
serve as the electron donor. Nitrogen, phosphorus, and sulfur will sometimes fulfill the role of electron
donor or acceptor, but are more often considered nutrients and are required in smaller proportions
than are carbon, hydrogen, and oxygen.
The specific growth rate of bacteria depends on the concentration of a carbon source (substrate) or the
nutrient that is most limiting. These growth kinetics are modeled by the Monod equation (Okpokwasili
and Nweke 2005), which shows how, at low substrate concentrations, the specific growth rate increases
directly with an increase in substrate concentration, while it levels out to approach a constant maximum
growth rate when substrate is plentiful. The specific growth rate of bacteria is particularly relevant in
bioremediation because bacterial growth rates (and proportionally, breakdown of the contaminant) will
likely slow as cleanup progresses, if the contaminant is the substrate.
1.4 Reduction and Oxidation Chemistry and Microbial Metabolism
Bacteria generate the energy they need to live by catalyzing (increasing, initiating, or transforming)
chemical reactions that transfer electrons from one molecule, known as the electron donor or
reductant, to another molecule, called the electron acceptor or oxidant. When the right electron donor
and acceptor are present, bacteria will consume them to grow and divide. The amount of energy
generated and available for bacterial growth by each reduction and oxidation or redox pair varies, and
each species of bacteria has enzymes to take advantage of only certain redox pairs. The contaminants of
concern may act as reductants or oxidants for in situ groundwater remediation.
The various terminal electron acceptors that exist naturally in groundwater are preferentially used and
exhausted in a specific order, according to their decreasing redox potential. In the environment, organic
matter in the aquifer matrix and groundwater plays the role of electron donor. The vast majority of
microbial metabolisms relevant to bioremediation use organic matter as an electron donor, and the
bacteria able to generate the most energy from it tend to dominate the microbial population. The
amount of energy released during electron transfer is controlled by the redox potential of the terminal
electron acceptor. There are a few important groups of bacteria that use inorganic reduced compounds
as a substrate. These microbes oxidize many of the same species reduced by anaerobic respiration and
Introduction to In Situ Bioremediation of Groundwater
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fermentation. Hydrogenotrophic methanogens are examples of bacteria that derive energy from
degrading inorganic compounds, since they oxidize hydrogen to water. Nitrifiers include aerobic bacteria
and Archaea that oxidize ammonia to nitrate, a process called nitrification that is extremely important to
nitrogen cycling. Sulfur-oxidizing bacteria oxidize sulfide to sulfur or sulfate, and iron-oxidizing bacteria
convert iron (II) to iron (III).
1.4.1 Aerobic Respiration
Aerobic bacteria use oxygen to oxidize organic molecules by removing electrons and converting the
organic molecules to carbon dioxide and water. Because of the high redox potential of oxygen, bacteria
able to use oxygen as a terminal electron acceptor will dominate wherever oxygen is present. Above
ground, aerobic environments are ubiquitous because they are in contact with the atmosphere, but
oxygen below ground surface can quickly be depleted by any aerobic microbial activity in groundwater.
1.4.2 Anaerobic Respiration and Fermentation
When oxygen is not present, bacteria commonly use nitrate, iron (III), manganese (IV), sulfate,
carbonate, or other available electron acceptors to oxidize organic matter, producing carbon dioxide and
other byproducts (Figure 4). Microbes exist that can use the contaminants for respiration for almost all
oxidized contaminants. Bacteria have been identified that use chemicals such as halogenated organic
compounds (such as PCE and trichloroethene [TCE]), selenium, arsenic, chromium (VI), technetium (VII),
and uranium (VI) as electron acceptors (Palmisano and Hazen 2003). This section discusses the electron
acceptors most commonly used by bacteria and most prevalent in the environment.
Nitrate is the first choice for electron acceptor after oxygen is depleted, and many aerobic bacteria
possess the enzymes to use nitrate to oxidize contaminants. Reduction of nitrate generates a sequence
of byproducts consisting of nitrite ions and the gases nitric oxide, nitrous oxide, and finally nitrogen. Use
of nitrate as electron acceptor is termed "denitrification" because it consumes nitrate.
Introduction to In Situ Bioremediation of Groundwater
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pE°(W)
; 0 Equivalents
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Figure 4. Dominant Terminal Electron-Accepting Process, Electron Acceptors, and Typical
Chemical Species Responses (Modified from AFCEE 2004, Bouwer and McCarty 1984)
Manganese and iron are often available for microbial use in the soil or groundwater. Iron-reducing
bacteria use iron (III) as an electron acceptor, reducing it to iron (II), or they can use manganese (IV),
reducing it to manganese (II). Once iron and manganese have been reduced, sulfate serves as an
electron acceptor and is converted by sulfur-reducing bacteria to sulfide, sulfite, or elemental sulfur.
When all external terminal electron acceptors have been exhausted, bacteria can use organic molecules
as both electron acceptors and donors in a metabolic pathway called fermentation. Fermentation
generates the least amount of energy because only a small fraction of the organic matter available can
be readily oxidized by microorganisms and because of the low redox potential of the reactions.
Fermentation can be divided into two categories: primary fermentation and secondary fermentation
(AFCEE 2004). Primary is the fermentation of substrates and amino acids to various volatile fatty acids,
alcohols, carbon dioxide, and hydrogen, while secondary is the fermentation of primary fermentation
products.
1.4.3 Direct Metabolism
Most bioremediation systems use a direct metabolic pathway, in which the contaminant of concern is
either an electron donor or acceptor, and the remedial system provides the presence of a
complementary oxidant or reductant and the right bacteria to take advantage of them. The growth rate
of bacteria depends on the concentration of substrate, which is the contaminant. As contaminants are
treated by a remedial system, contaminant concentrations may approach the minimum required for
bacterial growth, whether it be an acceptor or donor, and cause treatment to slow or stop.
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1.4.4 Cometabolism
Cometabolism is a term used to describe biological degradation from which bacteria do not derive any
energy. Bacteria secrete metabolic enzymes that break down complex organic matter around them for
easier digestion. These enzymes are often nonspecific and can operate on many different substrate
molecules, including those that the bacteria itself cannot use for energy. Enzymes such as methane
monooxygenase and ammonia monooxygenase are examples of enzymes that can oxidize a wide array
of substrates (Hazen 2009). Cometabolic treatment potentially can address even trace levels of the
contaminant, as long as the substrate the bacteria require for growth is maintained at acceptable
concentrations, because the bacteria do not rely on the contaminant for energy.
Cometabolism was once promoted as a method to treat TCE, but has rarely been used because the
intermediate epoxide produced inhibits biological activity. The TCE oxidation byproducts such as TCE
epoxide may result in the inactivation of the oxygenase activity caused by damage to the enzymes (Ely,
Hyman, and others, 1995). Inhibition and inactivation may be overcome by additional natural substrates
(Alvarez-Cohen and McCarty, 1991; Ely and others, 1997). Cometabolism may prove valuable for
treating other problematic contaminants such as A/-Nitrosodimethylamine (NDMA) and 1,4-dioxane
(Hatzinger and others 2008; Steffan 2007; Mahendra and Alvarez-Cohen 2006; Fournier and others
2009).
1.4.5 Abiotic Transformation
In some cases, the conditions created to encourage biological breakdown of contaminants will also be
conducive to abiotic chemical transformation of the contaminants, which occurs without the help of
organisms (Cwiertny and Scherer 2010). Added oxygen will oxidize many compounds without biological
catalysis, and hydrolysis of organic contaminants can happen spontaneously. Sulfide produced from
anaerobic sulfate reduction will precipitate some dissolved metal contaminants, such as lead, cadmium,
zinc, and copper (Lee 2003). Zero-valent iron can be added to support anaerobic bioremediation by
producing hydrogen as it oxidizes and to abiotically reduce contaminants.
The biological degradation pathway of 1,1,1-trichloroethane (TCA) generally stalls at chloroethane
(ATSDR 2006). Abiotic processes can play a key role in degradation of TCA to non-toxic end products.
TCA can be abiotically degraded to 1,1-dichloroethene (1,1-DCE) via a process called abiotic
dehydrochlorination (EPA 2009, ATSDR 2006). Figure 5 shows potential biotic and abiotic degradation
pathways for common chlorinated VOCs (CVOCs). At sites where TCA and TCE are co-contaminants, TCA
can inhibit Dehalococcoidesirom degrading TCE (Duhamel and others 2002). Abiotic
dehydrochlorination not only eliminates the inhibitory compound, but also creates a product that can be
degraded using the same bacteria and pathway as TCE. Alternatively, commercially available cultures
containing Dehalococcoides and Dehalobacter are capable of biologically degrading mixed plumes of
TCA and TCE.
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a
a A a a a
\ / dehydrochtorirrattoo v ,j'j*CI
c—c^ - •—^.-•F ^
Cl' CI
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a—c=c—H
CAC
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Figure 5. Example Biotic and Abiotic Degradation Pathways of Common CVOCs
(EPA 2009, O'Loughlin and Burris 2004)
Magnetite, biogenic iron-sulfide, and other naturally occurring minerals can also contribute to abiotic
transformation of chlorinated compounds (EPA 2009).
1.5 Conceptual Site Model
Adequate site characterization is critical to designing a successful remedy. The nature and extent of the
environmental impacts and the characteristics and interaction of the affected media need to be known.
Development of a conceptual site model (CSM) helps guide the characterization and subsequent design,
implementation, and performance of the remedy. The CSM will evolve through the life cycle of a project
as additional information is developed and generally includes a visual representation of the site (EPA
2011). At first, the CSM will consist of rough sketches that ideally evolve into a more comprehensive
representation of the available data, potentially using three-dimensional visualization and analysis
(3DVA), as discussed in Section 4.4. Figure 6 provides an example of a simplified pictorial CSM of an
exposure pathway analysis. The CSM is refined throughout the characterization and remediation process
at a site (EPA 2011).
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Poml
[Spill)
tmrinniatiBlBl -tttttum t»P«»r*.l!w
Figure 6. Conceptual Site Model - Exposure Pathway Schematic (ATSDR 2005)
Application of bioremediation is highly dependent on site characteristics, such as the aquifer type and
lithology. In particular, the magnitude and distribution of hydraulic conductivity affect the ability to
deliver amendments to the subsurface, where they are needed to maintain optimal conditions for the
targeted biological processes. Baseline characterization of the microbiology is essential to evaluate
whether the right microorganisms are present, if those microorganisms can be stimulated, and to
ascertain that no undesirable reactions will occur with the stimulants or daughter products. If
bioaugmentation is required or desired, the target treatment area must be properly conditioned to
support microbial growth. The following five sections summarize the key components of a CSM, with
emphasis on site characterization needed to assess and implement bioremediation.
1.5.1 Land Use and Risk
Important CSM components include past, current, and intended future land use and ecological value.
These factors will help to guide the site investigation, evaluate potential human health and ecological
risks, establish protective cleanup criteria, and evaluate acceptable control measures. For example, an
ISB system for a chlorinated solvent plume typically creates a reduced subsurface environment where
methanogenesis occurs naturally. In a residential neighborhood, the design would need to consider
vapor intrusion from contaminant vapors and potential methane gas as a critical design factor.
Conversely, indoor air quality may not be as important of a design factor in an industrial area dominated
by large manufacturing buildings with high air exchange rates.
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1.5.2 Geologic Setting
Understanding the geological setting and its heterogeneity is critical to developing a useful CSM.
Geological settings have unique characteristics that, if recognized early in the investigation, can help
with placement of soil borings and selection of analytical parameters. For example, some geological
settings are known for low baseline pH conditions that may adversely affect biological remedies. In
other geological settings, bedrock may have a significant influence on the direction of plume migration
through fractures, faults, and changes in porosity.
1.5.3 Hydraulic Properties of Contaminated Media
Proper characterization of hydraulic properties of the contaminated media may be one of the most
important components of a CSM. The following sections highlight some key hydraulic properties.
1.5.3.1 Hydraulic Conductivity
Hydraulic conductivity (K) is a measure of the ability of an aquifer matrix to transmit groundwater and is
important for designing delivery systems. It can be estimated in the field by slug tests, aquifer pumping
tests, and vertical hydraulic profiling. Hydraulic conductivities can differ by orders of magnitude within
an aquifer that may not be reflected in a single measurement of K. Slug tests are relatively affordable
and can be completed at several wells to help understand the distribution of K across a site, but can be
difficult to evaluate if significant differences are present between the gravel pack and aquifer materials.
Conversely, aquifer pumping tests tend to average K values as a result of their greater radius of
influence and may provide less detail regarding a specific location or specific depth interval. Vertical
hydraulic profiling provides high-resolution K data but cannot provide specific information regarding the
connectivity between higher conductivity zones in various profiles as an aquifer pumping test could. In
addition, vertical profiling provides an index of relative K at very small scale, versus measured K at larger
scale, which is very useful for understanding site heterogeneity. Often, a combination of aquifer test
methods is required to design an ISB system.
1.5.3.2 Porosity and Effective Porosity
Porosity (total porosity) is a measure of the void space in an aquifer. Specifically, total porosity is the
volume of the void space divided by the volume of aquifer matrix. Total porosity in bedrock aquifers is
described as primary and secondary. Primary porosity is the percentage of the voids in the rock at the
time of formation, and secondary porosity refers to the void space from fractures and dissolution (Fetter
2000). Primary and secondary porosity are important for estimating how much contaminant mass may
be present and can be useful for estimating amendment quantities. However, effective porosity is often
more important because it is a measure of the connected aquifer void space within the aquifer. Effective
porosity is lower than total porosity in most geological settings. When effective porosity is low,
amendment delivery systems may have difficulty treating the target area because of the poor
connections between the aquifer void space and fractures. A value for effective porosity is useful to
determine the following:
• The radius of influence of an injection well,
• The total number of injection wells, and
• Whether multiple screened intervals are required for the injection wells.
Introduction to In Situ Bioremediation of Groundwater 11
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Effective porosity can be estimated in the field by measuring groundwater flow velocity with a tracer
test (USGS 1999).
1.5.3.3 Groundwater Flow Direction and Velocity
Direction of groundwater flow is a key factor driving contaminant transport. Groundwater flows from
high to low hydraulic head. The hydraulic heads are often represented on a potentiometric surface map
(typically referred to as a water level map). Groundwater flow velocity, or seepage velocity, is a measure
of the groundwater flow rate through the aquifer pore space. Groundwater flow velocity is an important
parameter for selecting injection well placement and amendment quantities and type. For example,
sites with high groundwater flow velocities may require amendment to be added more frequently than
sites with low groundwater flow velocities.
High groundwater flow velocities can be incorporated into the amendment delivery design. For example,
injection wells can be used upgradient of inaccessible areas (such as under buildings) to deliver
amendments, allowing the natural and additional induced flow resulting from injection to transport the
amendments to the target treatment areas.
Preferential pathways exist in many geologic settings because of the heterogeneous distribution of more
transmissive zones caused by coarser-grained sediments or highly fractured bedrock. Groundwater flow
velocity through these zones will be higher than the average groundwater flow at the site. In addition,
groundwater flow velocities though preferential pathways can be exaggerated during the application of
amendments by the injection pressures. The increased groundwater flow velocities can cause
amendments to travel beyond the target application area or reach the ground surface (daylight).
1.5.3.4 Aquifer Matrix Diffusion Potential
Sedimentary aquifers commonly consist of heterogeneous layers or zones of different permeability and
transmissivity. Groundwater flows preferentially through more permeable zones as compared with the
less permeable zones. Bioremediation is more effective in the more permeable zones because liquid and
gas amendments infiltrate much more quickly through high permeability zones (Sale and others 2008).
Contamination often exists in the subsurface for many years before it is detected and remediated. This
delay allows the dissolved contaminants the time needed to diffuse from more permeable into less
permeable zones within the aquifer system. When remediation begins, the high permeability zones are
remediated more quickly and the concentration gradient between high and low permeability zones is
reversed. As a result, contaminants in the less permeable matrix will now diffuse back into the more
permeable matrix, causing contaminant levels in the more porous matrix to rebound after initial
treatment. The general matrix diffusion mechanism is shown in Figure 7. The less permeable matrix of
the aquifer becomes a new source area for contamination. Matrix back diffusion may persist for many
years after initial treatment.
Introduction to In Situ Bioremediation of Groundwater 12
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Advancing solvent plume Low permeability silts
Transmissive sand
Expanding diffusion halo in stagnant zone
Simultaneous inward and outward diffusion in stagnant zones
Figure 7. Conceptual Matrix Diffusion Mechanism (Adapted from NRC 2004)
Attainment of cleanup levels can be hindered by the slow release of contaminant mass from matrix
back-diffusion. Matrix back-diffusion can be significant where low-permeability zones are present within
the unconsolidated aquifer or where a dual porosity system exists as a function of adjacent lithologic
units having several orders of magnitude differences in K. Matrix back-diffusion can also play a
significant role in bedrock aquifers that exhibit sufficient primary porosity. For example, numerical
model simulations have demonstrated that back-diffusion from the matrix pore space (primary porosity)
to fractures (secondary porosity) will likely be the time-limiting factor in reaching groundwater cleanup
goals in some fractured bedrock environments (Lipson 2005). Matrix back-diffusion is observed much
more in sedimentary rocks than in igneous and metamorphic rocks.
Additional applications of amendments may be required to maintain a biologically active zone that will
ultimately reduce impacts to below the remedial action objectives. In some cases, several pore volumes
of treated groundwater may have to pass through the aquifer before objectives are met. Remediation of
aquifers where matrix back-diffusion is a factor may take longer and be more costly. Research is under
way to develop ways to estimate the rate of diffusion from the matrix into the groundwater.
1.5.4 Biogeochemistry
The geochemistry of a contaminated aquifer will control whether the necessary bacteria will grow in the
subsurface environment and what amendments are needed to help sustain the desired biological
processes. The following sections discuss some of the key geochemical parameters required to build a
CSM.
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1.5.4.1 Contaminant Types
Each contaminant type (for example, hydrocarbons, chlorinated aromatics, and aliphatics) will have an
effect on the site geochemistry. For example, excessive organic loading of an aquifer from landfill
leachate or a fuel release will result in biological activity that will readily consume oxygen and drive a
system to reducing conditions. The resulting reducing environment, in turn, may cause metals to
become mobile and create secondary groundwater impacts. Therefore, it is important to delineate the
contaminant plume and understand the effects of the contaminants on site geochemistry.
1.5.4.2 pH/Aquifer Buffering Capacity
The pH level in the subsurface is a significant factor for biological activity. The common range of pH for
most natural groundwater is between 5 and 8.5 (Kasenow 2010). Optimal ranges for dehalogenating
bacteria vary slightly in the literature, from 6 to 8 (AFCEE 2004) and 6.8 to 7.8 (Robinson and others
2009, Middeldorp and others 1999, Cope and Hughes 2001), for example. Sites that are well within the
generally recognized optimal pH ranges would pass the initial stages of the screening process for
potential biological treatment. However, biological treatment at sites on the margins or just outside the
optimal pH ranges should not be rejected until further site biological screening, such as bench testing, is
completed (AFCEE 2004). For example, local microbial populations at a site could have adapted to a low
pH environment and be able to sustain complete degradation of contaminants, or a low-pH tolerant
culture may be commercially available. In addition, pH buffering may be possible and cost effective.
Considerations of pH and buffering capacity for bioremediation are generally less about changing the
natural pH conditions — which can be a difficult endeavor for sustained and extended periods — and
are much more about establishing or maintaining the optimum microbial conditions after addition of
amendments and increased biological activity. Potential limiting pH conditions are common in anaerobic
bioremediation as a result of the generation of hydrogen through fermentation reactions and the
formation of organic acids that can exceed the buffering capacity of the aquifer. It is important to
measure the buffering capacity of the aquifer before a carbon source is added.
There are two laboratory approaches to measure buffering capacity. If an aquifer matrix is rich in
limestone and a high natural pH buffering capacity is anticipated, a laboratory acid titration test can be
completed on site soil and groundwater samples to determine the level of acid equivalents that will
reduce pH to levels outside optimal limits. This approach is referred to alkalinity testing. Alkalinity
testing results can be compared with stoichiometric calculations of the amount of electrons anticipated
to be liberated (acid to be produced) during the dechlorination process, given the site contaminant and
geochemical concentrations and the estimated donor quantities. If calculations indicate that more acid
may be produced than the aquifer has the capacity to buffer, a practitioner should consider the addition
of a buffering agent. Sodium bicarbonate is a typical pH buffering compound used, but it is a relatively
weak buffer and may be most appropriate for bioremediation applications where soluble substrates are
injected frequently. Stronger and more persistent buffering compounds such as magnesium hydroxide
or sodium phosphates may be used for bioremediation applications where slow-release substrates are
used (Henry 2010). The alkalinity testing will also provide an order of magnitude estimate of the amount
of buffering agent that will be needed to overcome aquifer acidity and maintain a near-neutral pH.
Introduction to In Situ Bioremediation of Groundwater 14
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Acidity testing is appropriate if a site is anticipated to have a low buffering capacity, or is in active
remediation with low pH conditions. Acidity testing consists of adding an alkali (such as sodium
bicarbonate or sodium hydroxide) to site soil and groundwater in a laboratory setting to determine the
equivalents of base needed to overcome aquifer acidity and maintain a near-neutral pH. Alkalinity and
acidity tests also provide insight on how the potential buffering requirements of the aquifer may affect
the feasibility of the bioremediation being planned.
The pH can also have other effects — besides the direct effect pH has on microorganisms — that could
negate the efficacy of bioremediation. For example, a decrease in pH can solubilize toxic metals that
were previously insoluble and create secondary environmental impacts. If an aquifer is known or
suspected to contain metals that may solubilize if pH is lowered, the practitioner may choose to add a
buffering agent to prevent pH from decreasing to a range that may solubilize metals, besides
considerations for microorganisms.
1.5.4.3 ORP
Oxidation-reduction potential (ORP) or redox potential describes the tendency of an aqueous solution to
either accept or donate electrons when a new species is introduced. Solutions with higher ORP are more
likely to oxidize new species, and solutions with lower ORP are more likely to reduce them. Figure 8
provides a summary of ORPs and the associated electron accepting process. A positive ORP is needed for
en
DC
=
~
1000
O/H2O
Denitrification: NO3/N2
Manganese Reduction: MnO2/Mn+2
Iron Reduction: Fe(OH)yFe+2
Alcohol Fermentation: CH2O/CH3OH
Sulfate Reduction: SO4"/H.;S
Methanogenesis: COj/CH4
Acetogenesis: CO2/CH3COOH (Acetic Acid)
H+/H2
Aerobic
(Oxygen as
Electron Acceptor)
580
i
H
I
V
Typical Prim ary
Substrates
(Electron Donors)
Figure 8. Estimated ORP of Commonly Monitored Species (ITRC 2005)
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aerobic oxidation of hydrocarbons and chlorinated solvents, while reductive dechlorination requires a
negative ORP, preferably below -200 millivolts (mV). Nitrate reducing conditions occur from 250 to 100
mV; reducing conditions for trivalent iron occur from 100 to 0 mV; reducing conditions for manganese
and sulfate occur from 0 to -200 mV; and methanogenesis occurs below -200 mV.
ORP can be a difficult parameter to measure accurately in the field. Down-hole probes or low-flow
pumps with flow-through cells are typically the most accurate methods of measuring ORP.
Although additives (substrates) can lower the ORP of a site to allow reductive dechlorination to occur,
the more oxidizing the natural conditions, the more substrate is needed, which could lead to other side
effects such as low pH and biological fouling (biofouling). Biofouling is attributed to the increase in
microbial populations and, perhaps more importantly, to the creation by cells of extracellular
polysaccharides. These slimy polysaccharides are important for the accumulation of microorganisms on
surfaces or within porous media and can contribute significantly to biofouling of a formation or injection
well. A portion of amendment goes to the creation of new bacteria (biomass). Eventually, continued
unchecked bacterial growth is likely to reduce circulation and injection of the amendment and may lead
to a plugged formation or injection well (ITRC 2002). Biofouling of injection or recirculation wells has
been observed at several sites because of the growth of biomass or biofilms with the well screen and
the surrounding sand pack. Several approaches have been used to mitigate these effects, and biofouling
should not be considered a major impediment to enhanced anaerobic bioremediation (AFCEE 2004).
1.5.4.4 Temperature
Each species of bacteria has an optimal range of temperature for growth. Growth rates increase with
temperature to an optimum near the top of the range and then quickly drop off as temperature
increases further. Bacteria are divided into groups based on their preferred temperature ranges:
psychrophiles are bacteria that grow best in temperatures below 20°C, mesophiles thrive between 25
and 35°C, and thermophiles prefer temperatures between 45 and 65°C.
Groundwater temperature varies geographically and seasonally and increases with depth. Figure 9
shows the average groundwater temperature across the continental United States. Shallow
groundwater can also be locally affected by precipitation as well as subsurface features such as process
equipment, utility lines, sewers, and other anthropogenic features. The biodegradation rate will slow as
the temperature drops, and many bacteria become inactive at temperatures less than 4°C. Therefore,
bioremediation in northern climates will be slower and may require additional design considerations.
For example, the optimal temperature for complete reductive dechlorination of PCE to ethene is
between 10 and 30°C. Below 10°C, the degradation half-lives of PCE and each of its daughter products
are substantially longer than at optimal temperatures (Dennis 2011). In colder environments, simple
soluble substrates (sugars or alcohol) may be more effective than more complex non-soluble substrates
(vegetable oil) because they are less viscous at lower temperatures and are easier to metabolize.
Circulation and subsequent heating of groundwater in a closed circuit could help maintain
biodegradation rates as well.
Introduction to In Situ Bioremediation of Groundwater 16
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Average Temperature of Shallow
Groundwater (°C)
Figure 9. Average Temperature of Shallow Groundwater across the continental
U.S. (EPA 2013, Ecosystems Research, Athens, GA)
1.5.4.5 Terminal Electron Acceptor Concentrations
Terminal electron acceptors are the typically native compounds used by organisms for respiration via an
electron transfer chain. Aerobes use oxygen as the terminal electron acceptor in the chain, and anaerobes
use various terminal electron acceptors. It is important to know the baseline concentrations at the site of
potential terminal electron acceptors such as dissolved oxygen, nitrate, manganese, iron, sulfate, and
carbon dioxide. These concentrations will describe the current redox state of the groundwater and what
quantity of amendments, if any, is necessary to eliminate the native electron acceptors. Native electron
acceptors compete with anaerobic dechlorination and must be reduced to a relatively narrow range in the
terminal electron accepting process for dechlorination to occur (AFCEE 2004).
If oxidative bioremediation is the targeted process, the presence of potential terminal electron
acceptors may mean less electron acceptor needs to be added to achieve complete biodegradation. If
reductive bioremediation is the selected remedial technology, competing electron acceptors need to be
used by bacteria and become depleted, causing bacteria to sequentially use the next available electron
acceptors in the process. Higher concentrations of native electron acceptors would generally indicate
that more electron donor will be required than is stoichiometrically demanded to degrade the
contaminant itself.
1.5.4.6 Nutrients or Growth Inhibitors
Nutrients are needed to sustain the growth of a bacterial population and include major nutrients nitrogen,
phosphorous, potassium, and minor nutrients sulfur, magnesium, calcium, manganese, iron, zinc, copper,
and trace elements. Although microbial activity could decrease if nutrients are not available in sufficient
amounts, nutrient deficiencies are typically not the growth-limiting factor when poor performance is
Introduction to In Situ Bioremediation of Groundwater
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observed. In fact, many practitioners do not add any nutrients beyond those that are naturally occurring.
Baseline characterization of an aquifer can assist in identifying potential nutrient needs, if any. If
necessary, nitrogen and phosphorous are usually added to the bioremediation system in a useable form
(such as ammonium for nitrogen and phosphate for phosphorous). However, nutrients can cause soil
plugging as a result of their reaction with minerals, such as iron and calcium, to form stable precipitates
that fill the pores in the soil and aquifer. Nutrients are required in larger proportions for aerobic systems,
compared with anaerobic systems, because of the higher growth rates for aerobic bacteria.
Some nutrients are competing electron acceptors in reductive dechlorination systems; therefore, the
amount and forms of nutrients require careful consideration. In some states (Michigan, for example),
groundwater antidegradation policies may limit or prohibit nutrient addition to aquifers without a
permit or restrict the use of certain compounds or product formulations. Many practitioners have opted
to add vitamins, primarily containing B-12, as a supplement at bioremediation sites. Vitamins have been
shown to increase rates of bioremediation (Environmental Security Technology Certification Program
[ESTCP] 2006). The addition of vitamins, however, is generally not required for bacterial activity to begin
and persist, but likely contributes to improved biological performance.
The presence of some compounds may slow or inhibit cell growth. For example, Dehalococcoides sp. has
been documented to be inhibited by hydrogen sulfide, chloroform, and 1,1,1-trichloroethane (He and
others 2005, Duhamel and others 2002). A good understanding of the target bioremediation bacterial
population is needed to identify what inhibitors to consider during the feasibility, bench-, or pilot-study
phases of a project.
1.5.4.7 Biostimulation and Bioaugmentation
Bioremediation is accomplished through exploitation of microbial metabolism. Biostimulation refers to
the addition of an electron donor (substrate) or electron acceptor, and bioaugmentation refers to the
addition of the bacteria that can break down the contaminant. It may be a challenge for native bacteria
to achieve required contaminant reductions without biostimulation or bioaugmentation for many types
of plumes, but it is possible given the right biogeochemical conditions. Naturally occurring
biodegradation is more common for contaminants that are degraded by aerobic bacteria (such as
gasoline products) in an environment where naturally occurring electron acceptors are common.
Significant contaminant reduction via natural biodegradation (without addition of amendments) by
anaerobic bacteria (such as chlorinated solvents) is less common. Although reduction of contaminants to
intermediate daughter products may occur, complete reduction of contaminants is less likely to occur at
sufficient rates to meet remedial objectives. Insufficient organic substrate is the most common limiting
factor.
Studies at chlorinated solvents sites with native target microbial populations have shown that
bioaugmentation test plots can outperform biostimulation test plots (Lendvay and others 2003).
Biostimulation was shown to take three to four times longer to achieve similar contaminant reductions.
Bioaugmentation is generally not needed at petroleum sites since the bacteria involved in hydrocarbon
bioremediation are ubiquitous in most environments. Conditions in a contaminated aquifer may not be
favorable for bacteria to thrive in the subsurface because of a number of potential reasons discussed in
Introduction to In Situ Bioremediation of Groundwater 18
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Section 1.6.4, but the presence of a target bacteria in an aquifer is a strong indication that
bioremediation is feasible. The absence of a target microbial population, however, does not preclude
application of bioremediation at a site. For example, Dehalococcoides may be present at a site, but at
population densities that are too low to detect and that become detectable only after amendments
have been added (biostimulation). The cost of amendments and their delivery to the target treatment
zones are often the highest portion of total project costs. Some practitioners consider bioaugmentation
in conjunction with amendment addition as a quicker means to obtain the required population densities
to reach complete reductive dechlorination. Although bioaugmentation is an additional project cost, its
use may reduce remediation time frames, as remedial goals are met more quickly. The net effect is a
low total project life cycle cost.
Molecular biological tools (MBTs) are becoming more widely available and cost effective for applications
in support of site characterization, remediation, and monitoring to determine microbial populations
within aquifers. Quantitative Polymerase Chain Reaction (qPCR) is the mostly commonly used method to
determine microbial populations. Other methods include microassays and Fluorescence In Situ
Hybridization (FISH) (ITRC 2013). Additional discussion regarding MBTs is included in Section 3.3.
1.5.5 Contaminant Distribution
The final component of the CSM described in this document is contaminant distribution. Contaminant
distribution is affected by each of the components of a CSM described in Section 1.6. A clear
understanding of the contaminant distribution and contaminant phases is critical for the proper design
of any remediation system. The contaminant mass distribution is a primary variable for bioremediation
sites used to calculate the quantity of amendment and identify the appropriate delivery method. The
lack of adequate characterization is one of the main reasons for poor remedial performance. Key
characteristics of contaminant distribution are discussed in this section.
1.5.5.1 SourceArea
According to Superfund guidance, " 'source material' is defined as material that includes or contains
hazardous substances, pollutants or contaminants that act as a reservoir for migration of contamination
to ground water, to surface water, to air, or acts as a source for direct exposure" (EPA 1991). The area
containing the source is usually where the release has occurred. Typical source areas are attributable to
underground and aboveground storage tanks (UST and ASTs), industrial lagoons, landfills, floor drains,
septic drainage fields, process equipment, chemical storage, and mine waste rock. The source area may
contain significant contaminant mass relative to the whole contaminated area, and impacts to the
vadose zone may be significant as well. A site may have multiple source areas. Biological approaches to
source area groundwater remediation have become more prevalent in recent years (CLAIRE SABRE
2010) and include the use of partitioning electron donors to try to attack NAPL source areas. Key design
factors are discussed in the field implementation section (Section 3).
1.5.5.2 Non-Aqueous Phase Liquid (NAPL)
Depending on the characteristics and amount of contaminant present, contamination may be
completely dissolved in the groundwater or exist as a NAPL, which is typically found within areas
considered source areas. NAPL co-exists with water in the pore space of an aquifer. Light non-aqueous
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phase liquids (LNAPL) tend to exist in the upper portion of the aquifer, while dense non-aqueous phase
liquids (DNAPL) tend to sink through the aquifer until they reach an impermeable formation. However,
more often than not, NAPL exists as isolated ganglia between pores in the form of residue rather than as
pockets of NAPL that fill all available pores (pooled NAPL) and are difficult to find and recover. The
presence of LNAPL is more readily apparent than DNAPL by direct observation of floating product in a
well, sheen on water during sampling, and coatings on sampling equipment. Often, the only clue to the
presence of DNAPL is if contaminant concentrations are at or near solubility limits or concentrations
rebound after some treatment takes place, as more DNAPL dissolves to equilibrate with the newly
treated water. However, contaminant rebound can also be attributed to other factors, including but not
limited to matrix back-diffusion or inflow of untreated groundwater.
Investigation using high-resolution site characterization (HRSC) strategies and technologies should be
considered in areas with potential NAPL, and more specifically DNAPL. The presence of NAPL can make
order of magnitude differences in the total aquifer contaminant mass. Conventional investigation
methods are more likely to miss DNAPL that may exist over small depth intervals in heterogeneous
geology. High resolution site characterization techniques are discussed in Section 4.3.
Over the past several years, the application of ISB to treat DNAPL source areas has become more
common. It has been demonstrated that dechlorinating organisms can tolerate concentrations of
chlorinated ethenes near the solubility limit (ITRC 2008). Biological degradation occurs only in the
dissolved phase, but other mechanisms accelerate source zone mass removal, as stated in ITRC 2008:
• Increasing the concentration gradient at the DNAPL-water interface, which increases the rate of
DNAPL dissolution;
• Partially biodegrading parent compounds near the DNAPL-water interface, producing less-
chlorinated daughter products (c/s-l,2-dichloroethene cis-l,2-DCE) and vinyl chloride [VC]) that
are more mobile in groundwater than TCE and PCE; and
• Under some conditions, the electron donor solution or its degradation products abiotically
enhance DNAPL mass transfer rates through cosolvency, desorption, or dissolved organic matter
or surfactant partitioning. Studies are currently under way to demonstrate and validate the
application of this approach (ESTCP 2013)
For instance, some bacteria produce natural surfactants that help the bacteria break down the NAPL at
its surface interface (Banat and others 2000). Bench and pilot studies have demonstrated that the
application of bioremediation in DNAPL source areas is a feasible remediation technology capable of
reducing contaminant concentrations in groundwater within the source area and enhancing the removal
of non-aqueous and sorbed contaminant mass (Hood and others 2008).
1.5.5.3 Dissolved Plume
The dissolved plume is located in and downgradient of the source area. The shape, concentration, and
vertical and horizontal extents are controlled by components of the CSM. The most important factors
are the type of contaminant; the initial concentration; and the rates of advection, dispersion, and
diffusion. Biological approaches to remediation of dissolved plumes in groundwater have been widely
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applied across the United States at aerobic sites since the 1980s and at anaerobic sites since the late
1990s. Important design factors to consider are discussed in the field implementation section
(Section 3).
1.5.5.4 Lateral Extent, Thickness, and Depth
The majority of the costs associated with bioremediation are attributable to the quantity of
amendments required and the methods needed to deliver them to targeted treatment areas. The lateral
extent, the thickness of the affected zone, and the total depth required to reach the contaminated areas
strongly influence the selection of a remedial approach. Target treatment areas that are limited in
horizontal and vertical extent with low concentrations and limited potential for rebound may be ideal
for direct injection and will have relatively low implementation costs. Conversely, target treatment areas
that are expansive in horizontal and vertical extent with high concentrations and a high potential for
rebound will likely require permanent injection wells with multiple screen intervals, multiple
amendment applications, and will have significantly higher implementation costs.
Complex and highly heterogeneous sites often involve several target treatment zones. Each treatment
zone may require unique delivery and amendment designs, depending on the differences in
hydrogeology, depth, or co-contaminants. As a result, a successful pilot- or bench-scale study directed at
one zone does not guarantee success for the other zones. Sites with dissimilar target treatment zones
could become more expensive to treat than expected if this level of detail is not addressed at the site
characterization and feasibility study stages of a project.
1.5.5.5 Contaminant Mass Flux and Mass Discharge
The final key characteristics of contaminant distribution are mass flux and mass discharge. Mass flux is
the flow rate of contaminant mass through a defined area, usually a portion of a plume cross section.
Mass flux is expressed as mass per time per area. Mass discharge is the integration of mass flux
measured across an entire plume and thus represents the total mass of any contaminant plume
conveyed by groundwater through a defined plane. Mass discharge is expressed as mass per time. In
addition to defining the source strength and plume attenuation rate, mass flux estimates can identify
areas of a plane where most of the contaminant mass is moving. Mass flux and mass discharge can be
measured using transect methods, where concentration and flow data are collected from new or
existing monitoring points and integrated; well capture and pump tests, where groundwater is extracted
from wells while flow and mass discharge are measured; and passive flux meters, which are instruments
that estimate mass flux directly within wells (ITRC 2010).
Incorporating mass discharge information into the CSM will help improve remediation efficiency and
shorten cleanup time, particularly at sites with multiple source areas or where plumes cross multiple
stratigraphic units. Generally, the majority of contaminant mass flows through a small portion of a cross-
sectional area of an aquifer. Guilbeault and others (2005) studied three sites in North America using
cross-sectional transects that 75% of contaminant mass discharge occurs through 5% to 10% of the
plume cross-sectional area. Mass flux and mass discharge are extremely useful parameters to consider
in designing an amendment delivery system, though the cost to collect the data needed to calculate
mass flux and mass discharge increases with desired accuracy. The added costs may be justified if there
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is a possibility that the target treatment area could be reduced or more accurately located and
addressed. More targeted treatment can reduce costs and lead to more effective remediation.
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2.0 STRATEGIES FOR GROUNDWATER BIOREMEDIATION
There are a wide array of groundwater bioremediation strategies available, each appropriate for specific
contaminants and site conditions. The following sections explain the four main categories of
bioremediation strategies (aerobic, anaerobic oxidative, anaerobic reductive, and cometabolic), the
contaminants these strategies can treat, key microbe summary, summary of electron donors and
acceptors, and general implementation approaches.
2.1 Aerobic Bioremediation
Aerobic bioremediation takes place in the presence of oxygen, which is the electron acceptor. With few
exceptions, it relies on the direct microbial metabolic oxidation of a contaminant. The primary concern
when an aerobic bioremediation system is designed is delivery of oxygen. Aerobic bioremediation
technologies have been used at Superfund sites for more than 20 years.
2.1.1 Common Applicable Contaminants
Aerobic bioremediation is most effective in reducing non-halogenated organic compounds to carbon
dioxide and water. Typically, aerobic bioremediation is applied to treat BTEX and diesel and jet fuel
releases, often from USTS or ASTs (Farhadian and others 2008). Heavier hydrocarbons (those with
higher molecular weight), such as lubricating oils, generally take longer to biodegrade than lighter
products, but bioremediation can also be feasible for heavier fuels. In addition, although much less
common, aerobic bioremediation has also been successfully applied to treat other solvents such as
acetone, non-halogenated SVOCs including alkenes and alkanes found in fuels, some polycyclic aromatic
hydrocarbons (PAHs), and pesticides and herbicides. For example, 11 Superfund projects in Appendix A
selected aerobic treatment to treat the pesticide pentachlorophenol (PCP). Eight of these 11 projects
also involve PAHs or naphthalene contamination. Project performance may be available at the relevant
websites for these Superfund sites.
Aerobic bioremediation is more successful for simpler PAH compounds such as naphthalene. Biologically
treating more complex cyclic compounds, such as benzo(a)pyrene, is considerably more difficult.
Bioremediation of SVOCs and PAHs is far more common in soil remediation applications because these
types of contaminants are more likely to be sorbed to soils than dissolved in water. However,
bioremediation of groundwater containing select compounds from these groups has been documented
(see for example Brubaker and others 1992). Aerobic bioremediation has also been used for a wide
range of other contaminants, including vinyl chloride, DCE, methyl tert-butyl ether (MTBE),
chlorobenzenes, ketones, some pesticides (such as 2,4-Dichlorophenoxyacetic acid), and some
nitroaromatics such as dinitrotoluene.
2.1.2 Key Microbe Summary
Many reduced contaminants can be aerobically degraded by aerobic bacteria already present in the
subsurface environment. Many species can metabolize the less recalcitrant organic contaminants
because most aerobic heterotrophic bacteria can make use of a range of substrates. Some of the most
common aerobic bacteria with the ability to degrade BTEX and PAHs, among other common
contaminants discussed in Section 2.1.1 are Pseudomonas, Alcaligenes, Sphingomonas, Rhodococcus,
and Mycobacterium. These microbes have been well documented to degrade pesticides and
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hydrocarbons, both alkanes and polyaromatic compounds (Vidali 2001, Hendrickx and others 2005,
Bamforth and Singleton 2005) and are generally considered ubiquitously distributed in the natural
environment (Bamforth and Singleton 2005). Fluorescent Pseudomonas strains, especially P. putida
strains, are often isolated as BTEX degraders from BTEX- and gasoline-contaminated sites (Hendrickx and
others 2005).
2.1.3 Sources of Electron Acceptor
Oxygen is the electron acceptor required for aerobic bioremediation. Oxygen can be added directly to
the subsurface or oxygen-releasing compounds can be applied, which release oxygen as they dissolve or
decompose. Common oxygen-releasing chemicals are calcium and magnesium peroxides, hydrogen
peroxide, and ozone.
Calcium peroxide and magnesium peroxide break down in water to their hydroxide forms, releasing
hydrogen peroxide. Hydrogen peroxide breaks down further in water, completely decomposing into
oxygen and water within 4 hours (EPA 2004). Hydrogen peroxide is generally toxic to microorganisms at
concentrations above 100 parts per million (ppm), though microbes can tolerate up to 1,000 ppm
hydrogen peroxide with proper acclimation. Ozone also decomposes into oxygen in water and is 10
times more soluble in water than oxygen itself (EPA 2004), though it is also toxic to microorganisms at
higher concentrations.
Of these chemical amendments, magnesium peroxide and ozone provide the highest relative oxygen
delivery efficiency, but magnesium peroxide is a significantly longer-term oxygen-releasing chemical
than ozone. Many of these compounds are used in high concentrations as chemical oxidants. Residual
oxygen may not have the oxidative power to continue chemical oxidation, but may be sufficient to
support biological activity. As a result, it is common to transition to aerobic bioremediation after
chemical oxidation remedies are completed.
2.1.4 Delivery Mechanisms
Oxygen and oxygen-releasing compounds can be delivered to the groundwater via several methods,
depending on their physical properties, site hydrogeology, and the desired delivery efficiency.
2.1.4.1 Gas-Ph ase Delivery
Injection of gas into groundwater to stimulate or enhance aerobic biodegradation is called biosparging.
The efficacy of biosparging depends primarily on the permeability of the aquifer and the
biodegradability of the contaminant (EPA 2004). Intrinsic permeability (related to effective porosity) is a
measure of the ability of soil to transmit fluids and is the single most important characteristic of the soil
in determining the effectiveness of biosparging because it controls how well oxygen can be delivered to
the subsurface microorganisms (EPA 2004). Treatment in zones with low permeability will be limited by
diffusion. Air, oxygen, and ozone can all be delivered using biosparging.
Biosparging differs from air sparging in that air sparge systems are designed to remove contaminants
through volatilization and require a soil vapor extraction system to capture volatilized gases, while
biosparging delivers air at lower flow rates as well as nutrients, if needed, to stimulate biodegradation
and minimize volatilization. Biosparging has been selected for five projects included in the Appendix A
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data set. Some degree of volatilization of contaminants will occur with biosparging (EPA 2004), which
has the potential to build up pressure or cause hazardous atmospheres. When biosparging is applied in
potentially sensitive areas with basements, sewers, or subsurface confined spaces, biosparging can be
combined with soil vapor extraction (EPA 2004) and sub-slab depressurization systems to control vapor-
phase contaminants.
An alternative method of dissolving oxygen gas is using gas diffusers. Gas diffusers consist of a cartridge
containing a semi-permeable membrane designed to be submerged into a groundwater well and
pressurized with oxygen. The semi-permeable membrane allows oxygen molecules to pass through into
the liquid, and the increased pressure of oxygen causes the water to become supersaturated compared
with oxygen concentrations possible under atmospheric conditions. This method can be used to dissolve
air or oxygen in groundwater more efficiently than biosparging, because no gas is lost from bubbling up
out of the groundwater. Ozone, however, is too reactive to use with the delicate membranes of the
current gas diffusers on the market, and it must be applied to the subsurface in a manner similar to
biosparging.
2.1.4.2 Liquid- and Solid-Phase Delivery
Liquid delivery of oxygen to the subsurface can be achieved in several ways to support aerobic
bioremediation. The most direct is injection of water supersaturated with oxygen. This method uses a
technology similar to that used for gas diffusion discussed in the previous section, but the mechanical
infusion of oxygen into the water occurs before it is applied to injection wells. Ozone is also commonly
applied as a solution to support aerobic bioremediation. Ozone provides an oxygen delivery efficiency
that is higher than other chemical amendments, but lower than biosparging. Hydrogen peroxide can
deliver significant oxygen to a saturated zone; however, hydrogen peroxide decomposes and liberates
oxygen faster than the oxygen can be biologically used (EPA 1990).
Calcium and magnesium peroxide can be injected into the saturated zone as a solid or in slurry form
(EPA 2004). Magnesium peroxide is more commonly used because it dissolves more slowly, prolonging
the release of oxygen. In their solid forms, these chemicals can be mixed with water or in slurry for
injection. Solids that can be injected are generally fine grained (able to pass through a 0.02-inch
opening). Solids in water applications usually require greater and continuous agitation of the mixed
product to ensure application of a consistent, homogenized mixture. Lower injection pressures (below
100 pounds per square inch [psi]) are typically adequate for delivery of water-based mixtures. The mixed
material can be applied using injection wells with openings of adequate size to allow solids to pass
through or applied by direct injection using specialized drilling tooling (for example, Geoprobe Systems
Pressure Activated Injection Probe).
Slurries are typically mixed on site and then injected soon after mixing to minimize settling of the
product. Larger-volume injections may require intermittent mixing of batches to prevent settling of the
oxygen-releasing solids and maintain an even distribution from injection to injection. Slurries are
primarily applied via direct injection using the tooling previously mentioned. Slurries are typically
injected under higher pressures (100 to 500 psi), depending largely on the receiving material. These
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higher injection pressures will fracture soils, in most cases, which need to be considered given specific
site conditions.
2.1.5 Common Byproducts
The by-products of aerobic bioremediation are generally carbon dioxide and water. Excessive calcium,
magnesium, or iron in groundwater can react with carbon dioxide. The products of these reactions can
adversely affect the operation of an ISB system. Crystalline precipitates or "scale" is formed when
calcium, magnesium, or iron reacts with phosphate or carbon dioxide. Scale can constrict flow channels
and can also damage equipment, such as injection wells and sparge points.
The precipitation of calcium or magnesium phosphates can also tie up phosphorus compounds, making
them unavailable to microorganisms for use as nutrients. Precipitation of calcium or magnesium
phosphates can be minimized by using tripolyphosphates to act as sequestering agents to keep the
magnesium and calcium in solution (prevent the metal ions from precipitating and forming scale) (EPA
2004).
When oxygen is introduced to the subsurface as a terminal electron acceptor, it can react with dissolved
iron [Fe(ll)] to form an insoluble iron precipitate, ferric oxide. The precipitate can be deposited in aquifer
flow channels, reducing permeability. The effects of iron precipitation tend to be most noticeable
around injection wells, where the oxygen concentration in groundwater is highest, and can render
injection wells inoperable. Lower injection rates and higher pressures are often indicators of a decrease
in injection well performance. Routine injection well maintenance may be required.
At least one aerobic pathway is available for all of the BTEX compounds that include degradation to
catechol or a substituted catechol. The byproducts of BTEX metabolism are not considered
contaminants of concern.
2.2 Anaerobic Oxidative Bioremediation
Anaerobic oxidative bioremediation, like aerobic bioremediation, also relies on the direct microbial
metabolic oxidation of a contaminant and is an alternative to aerobic bioremediation in anaerobic
aquifers. Generally, aerobic conditions allow for a higher rate of biodegradation of reduced
contaminants than anaerobic conditions. As a result, remediation strategies often introduce oxygen to
anaerobic environments in an attempt to employ more efficient aerobic microbial processes. However,
the overall oxygen demand from dissolved metals such as iron and manganese is often overlooked and
underestimated. Even if oxygen demand is accounted for, the result of oxygen delivery may interfere
with the injection infrastructure. Oxygen will readily react with dissolved iron(ll) to form an insoluble
iron(lll) precipitate, which decreases the permeability of the aquifer and may foul injection tools and
wells. Therefore, it is advantageous to promote anaerobic oxidative bioremediation where oxygen levels
are already depleted, an appropriate metabolic pathway exists for the target contaminants, and other
conditions are conducive to this approach (as discussed below).
The key concern when an anaerobic oxidative bioremediation system is designed is the availability of a
carbon source, nutrients, and an electron acceptor. The rate of degradation is typically limited by the
availability of an electron acceptor; however, carbon or nutrient amendments may be necessary as well.
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In the absence of oxygen, anaerobes will preferentially use alternative electron acceptors based on the
amount of free energy gained from reduction of a given electron acceptor, as described above. The
preferential electron acceptor in the absence of oxygen is nitrate, followed by manganese(IV), iron(lll),
sulfate, and finally carbon dioxide.
2.2.1 Common Applicable Contaminants
Several contaminants can be anaerobically oxidized, including aromatic hydrocarbons, fuels, and some
chloroethenes. Aromatic hydrocarbons associated with petroleum and fuel releases such as BTEX will
undergo anaerobic oxidative biodegradation. Remediation product suppliers provide sulfate enhanced
amendments to promote anaerobic oxidative bioremediation at petroleum-contaminated sites. Toluene
and xylenes are more readily oxidized anaerobically than are benzene and ethylbenzene; however,
degradation of benzene and ethylbenzene has been documented in manganese and iron reducing
environments (Villatoro-Monzon 2003). Naphthalene has also been observed to degrade anaerobically
by sulfate-reducing bacteria (Meckenstock 2000). Additional aromatic hydrocarbons that are degraded
anaerobically include phenol, cresol, and benzoic acids (Suflita 1991).
Biodegradation of chloroethenes is typically associated with reductive dechlorination processes.
However, DCE and vinyl chloride appear to be anaerobically oxidized in iron(lll)-reducing and
methanogenic conditions (Bradley 2007). Vinyl chloride may also be oxidized in sulfate-reducing and
humic acid-reducing environments (Bradley 1997). However, even under nominally anaerobic
conditions, very low levels of oxygen (much less than the typical reporting limit of 1 milligrams per liter
[mg/L]) may support aerobic biodegradation of chloroethenes, potentially confounding the
interpretation of results from laboratory and field tests designed to stimulate anaerobic oxidation
(Gossett 2010).
2.2.2 Key Microbe Summary
Several microbes involved in bioremediation can adapt to aerobic and anaerobic conditions. These
microbes are called facultative, and while active in both aerobic and anaerobic environments, facultative
microbes degrade contaminants at a slower rate in the absence of oxygen. Microbes that use nitrate as
an electron acceptor tend to be facultative (Firestone 1982). Facultative microbial action accelerates the
depletion of nitrate because it is used as a nutrient as well as an electron acceptor.
Strict anaerobes will be active only in reduced environments and will use electron acceptors such as
sulfate or carbon dioxide. Sulfate-reducing bacteria are obligate anaerobes (they require an anaerobic
environment to thrive). Desulfovibrio is the most well studied sulfate reducer.
2.2.3 Sources of Electron Acceptor
Several inorganic compounds commonly found in aquifers may act as electron acceptors for anaerobic
oxidative bioremediation. Various commercial products are available that can supply electron acceptors
to drive the anaerobic oxidation process. These products most commonly contain iron(lll), nitrate, or
sulfate. The selection of an electron acceptor or product will depend on the contaminant and the
optimal oxidation-reduction state of the targeted bioremediation process. For example, a product
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containing primarily iron(lll) should be used if degradation of toluene is observed at a site that is mildly
anaerobic and data indicate that the oxidation-reduction state is iron-reducing.
Nitrate is highly soluble in water and, after oxygen, provides the most free energy for microbial action.
Nitrate is also mobile in an aquifer. However, nitrate concentrations in groundwater above 10 mg/L
have negative toxicological effects on humans and animals. Therefore, care must be taken if
groundwater is to be amended with nitrate.
Iron(lll) salts are only slightly soluble in water, but when used as an electron acceptor, iron(lll) is reduced
to iron(ll), which is much more soluble in water. Iron(lll) has a particularly low electron accepting
capacity for its mass, and therefore iron(ll) may quickly exceed water quality thresholds in groundwater
as it reacts and dissolves.
Sulfate is very soluble in water, will not sorb appreciably, and is generally unreactive. Sulfide, the end
product of sulfate reduction, precipitates with iron(ll) and is effectively immobilized. However, in acidic
environments, sulfide can produce hydrogen sulfide gas, which is toxic to breathe.
2.2.4 Delivery Mechanisms
Since the most common electron acceptors used for anaerobic oxidative bioremediation are soluble in
water, the products are typically delivered to the subsurface in a solution via injection wells or direct
injection using drilling tooling similar to the delivery methods discussed in Section 2.1.4.2. The key to
successful delivery is matching the proper concentration of solution to the remedial objective and
avoiding potential negative impacts to the groundwater.
2.2.5 Common Byproducts
Anaerobic oxidative bioremediation may produce metabolic byproducts that can be problematic when
oxygen is not the terminal electron acceptor, and if certain compounds exist in an aquifer. The most
common of the issues discussed in this section that practitioners have observed is mobilization of
metals. Arsenic mobilization is of particular concern in areas with naturally occurring arsenic in soils or
bedrock. Mobilized metals will persist until the oxidation-reduction state shifts back to oxidative and
metals form oxides, making them generally immobile. The shift back to oxidative conditions occurs
naturally downgradient of a biological treatment area; however, an engineered approach such as an air-
sparge wall could be installed to induce oxidative conditions if receptors are present.
A few additional examples of problematic byproducts are as follows:
• Nitrate is reduced to nitrite, nitric oxide (reactive intermediate), nitrous oxide (reactive
intermediate), nitrogen gas, or a combination of these byproducts, depending on the microbes
that are present. The latter three are gaseous byproducts that can dissolve into groundwater to
some extent, but will generally escape into the vadose zone; however, the gaseous byproducts
may become trapped within pore spaces, displacing water and reducing the hydraulic
conductivity of the saturated matrix.
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• Manganese (IV) and iron (III) are reduced to soluble manganese (II) and iron (II). These
dissolved-phase metals may contribute to secondary groundwater plumes and elevated total
dissolved solids.
• Sulfate is reduced to sulfite and sulfide. The end product of sulfate reduction is sulfide. If there
are not enough dissolved metals to precipitate the sulfide, hydrogen sulfide gas is generated,
which is toxic and flammable and could result in vapor intrusion issues given the depth of the
plume and characteristics of any overlying buildings.
• Fermentation generates hydrogen ions, which can lower the pH of the groundwater to levels
where the key bacteria cannot survive. In addition, carbon dioxide is reduced to methane, which
can support a community of microbes called methanotrophs but could result in vapor intrusion
issues.
2.3 Anaerobic Reductive Bio remediation
Anaerobic reductive bioremediation takes place in the absence of oxygen. It relies on the presence of
biologically available organic carbon naturally or the application of a reduced carbon source, also
commonly called organic substrate, into groundwater to create and sustain anaerobic conditions and
the bioreduction of contaminants, such as chlorinated solvents (EPA 2001a), by generating hydrogen
through fermentation reactions. Because chlorinated solvents exist in an oxidized state, they are
generally much less susceptible to aerobic oxidation processes. However, they are susceptible to
microbial reduction under anaerobic conditions. The key concerns when an anaerobic reductive
bioremediation system is designed is the competition of native electron acceptors (such as oxygen,
nitrate, iron, and iron sulfate) with the contaminant, the presence of bacteria capable of completely
reducing contaminants, and the effective delivery of the substrate to all portions of the aquifer that is
contaminated.
Injected organic substrates are first fermented to hydrogen and low-molecular weight fatty acids, which
in turn provide a source of carbon and energy to the microorganisms. Microorganisms will consume
competing native electron acceptors beginning with the most oxidized sequentially to the least oxidized.
Once native electron acceptors have been eliminated or depleted, target contaminants will be the most
efficient electron acceptors. In many cases, microorganisms use the highly oxidized contaminants in a
respiratory mechanism and are able to derive metabolically useful energy (EPA 2000 and AFCEE 2004),
requiring either continuous or intermittent substrate replenishment to maintain favorable conditions for
the microorganisms. Other processes such as anaerobic cometabolism could also occur in the
subsurface as a result of injection of carbon substrates and the creation of reducing conditions.
Anaerobic reductive bioremediation can cost-effectively remediate contaminated sites if the site can be
engineered to provide appropriate growth conditions favorable to native contaminant degrading
microbes or commercially available microbial cultures, and if contaminants are susceptible to reductive
bioremediation (such as chlorinated solvents or perchlorate). Although anaerobic bioremediation has
been applied at hundreds of sites to date, many sites have not been closed using anaerobic reductive
bioremediation alone. Many factors — including source area mass, the presence of NAPL, aquifer
characteristics, and cleanup objectives — will dictate whether other technologies may need to be
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implemented before, in concert with, or after implementation of anaerobic reductive bioremediation.
ITRC 2011 presents an in-depth discussion of developing an integrated site strategy to account for these
factors.
2.3.1 Common Applicable Contaminants
Contaminants that can be degraded using anaerobic reductive bioremediation include halogenated VOCs,
munitions, some dissolved metals, perchlorate, and nitrate. Given the increased interest in anaerobic
reductive bioremediation, more detail is given in the following sections on these applications. Enhanced
reductive dechlorination (ERD) can be used for chlorobenzenes, chlorinated pesticides, and chlorinated
SVOCs, but these compounds are more difficult to degrade than chlorinated ethenes (ESTCP 2008).
2.3.1.1 Halogenated VOCs
Anaerobic bioremediation may be used to degrade chlorinated contaminants, such as biodegradation of
PCE and TCE to ethene, 1,1,1-TCAto ethane (with further degradation to other non-toxic compounds)
and carbon tetrachloride (CT) to methane.
The most common halogenated VOCs include PCE, TCE, 1,2-dichloroethane (DCA), and CT that are called
chlorinated aliphatic hydrocarbons (CAM). Generally, the more chlorinated the CAM, the more
appropriate it is to use anaerobic versus aerobic degradation processes. However, less chlorinated
compounds and dechlorination products such as DCE, VC, and chloroethane can be degraded using
either anaerobic or aerobic bioremediation technologies (ESTCP 2008). As seen in Figure 2, halogenated
VOCs are the contaminants most commonly treated by in situ groundwater bioremediation projects in
the EPA's Superfund program.
At some sites, anaerobic reductive bioremediation of PCE and TCE may undergo incomplete degradation
(stalls) to DCE or VC. Many factors can cause stall, such as:
• Reductive dechlorination is most effective and efficient under sulfate-reducing to methanogenic
conditions. The inability to achieve these negative redox conditions can cause a stall at DCE.
• Microorganisms generally gain more energy from dechlorination of more highly chlorinated
CAHs (such as PCE and TCE). Dechlorination of daughter products (DCE and VC) may not proceed
until parent products are sufficiently depleted (AFCEE 2004).
• Dehalococcoides mccartyi is the only known bacteria that can achieve complete dechlorination
of chlorinated ethenes (Ernst 2009; Loffler and others 2013). Sometimes this bacterium is not
present at population densities required to sustain complete dechlorination.
• Co-contaminants or other geochemical conditions such as pH can inhibit the microbial
population, such as inadequate electron donor availability or unfavorable geochemistry.
Remediating chlorinated VOCs can be challenging, and the degree of success is subject to hydrogeological
and biogeochemical conditions. Chlorinated VOCs can be remediated in situ through anaerobic reductive
bioremediation, but it is not appropriate for every site. In addition to limitations from the characteristics of
the aquifer that can make it difficult to access the contamination, the timeframe required for complete
dechlorination can be months to years, depending on groundwater flow velocity and matrix diffusion. In
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addition, native microbial populations can compete with dechlorinating microbes, or other conditions can
exist as listed above, that can lead to incomplete degradation pathways. The bioremediation process can
change groundwater pH, and redox conditions, such that the solubility of some metals increases and cause
secondary water quality impacts. Some of these limitations also apply to other remedial techniques and
are not unique to bioremediation (AFCEE 2004).
2.3.1.2 Nitrate and Sulfates
Nitrate is essential for plant growth, but is potentially toxic to human and animal life at moderate
concentrations. Sources of nitrate in groundwater include atmospheric deposition from fossil fuel
burning, runoff from fertilizer use, leaching from animal wastes from confined animal feedlot operations
and dairies, septic tanks and sewage, landfills, and erosion of natural deposits. Impacts to groundwater
from sulfate are derived from some sources similar to nitrate sources. Two common sources are
agricultural sulfate and geochemistry changes to aquifers that contain sulfide minerals.
Anaerobic reductive bioremediation of nitrate- and sulfate-contaminated groundwater can be achieved
by biostimulation of native nitrate-reducing and sulfate-reducing microbial communities. Anaerobic
bioremediation was selected to successfully treat nitrates at two projects listed in Appendix A. The
conditions that facilitate the process are part of the aquifer reduction and competing electron acceptor
elimination step for establishing reductive dechlorination.
2.3.1.3 Perchlorate
Perchlorate is both a naturally occurring and man-made chemical that is used to produce rocket fuel,
fireworks, flares, and explosives. Perchlorate-contaminated groundwater can be treated using anaerobic
reductive bioremediation or aboveground bioreactors. Perchlorate is degraded by a three-step
reduction process where perchlorate is reduced to chlorate, chlorite, and finally to chloride and oxygen.
Perchlorate-reducing microorganisms are generally ubiquitous in the environment, but bioaugmentation
is sometimes needed to reach population densities required for treatment (ITRC 2005). These microbes
produce an enzyme that allows them to lower the perchlorate activation energy for reduction and use
the perchlorate as an electron acceptor. Once the aquifer has been conditioned with an electron donor
(organic substrate) to eliminate the primary competing electron acceptors — oxygen and nitrate —
perchlorate reduction can occur. The presence of molybdenum might be required by the
microorganisms (ESTCP 2008; Stroo and Ward 2009). Anaerobic reductive bioremediation has been
selected to treat perchlorate or Royal Demolition Explosive (RDX) at three projects listed in Appendix A.
2.3.1.4 Pesticides and Herbicides
Some pesticides and herbicides can be treated using anaerobic reductive bioremediation.
Organochlorine pesticides, such as toxaphene and dieldrin, are persistent in the environment and
adsorb strongly to soils but do not decompose naturally at a significant rate. Field studies have shown
that the application of solid organic carbon and zero-valent iron to groundwater can effectively treat
organochlorine pesticides. These amendments create a reduced environment that supports relatively
rapid and complete dechlorination of many chlorinated compounds (Seech 2008).
Introduction to In Situ Bioremediation of Groundwater 31
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2.3.1.5 SVOCs and PCPs
SVOCs from wood treating wastes include creosote and PCP, which are typically treated using
bioremediation in the soil matrix, in sediments, or in mulch. In groundwater, typical treatment includes
anaerobic reductive dechlorination followed by anaerobic oxidation. However, there are examples of
full-scale in situ implementation that report success using a carbon source supplemented with oxygen
(Fields 2010). Since PAHs are the main components in creosote, bioremediation technologies for PAHs
can also be considered for creosote contamination (Zhang 2010), although PAHs above three rings are
recalcitrant to bioremediation under any scenario because of their typically low solubility.
2.3.1.6 Dissolved Metals
Metals cannot be destroyed through bioremediation technologies. Rather, microbes can remove
dissolved metals from solution by reducing them to a more insoluble valence state. Immobilization
reduces the mobility of contaminants by altering the physical or chemical characteristics of the
contaminant, causing it to precipitate out of solution or to sorb onto the soil. Furthermore,
microorganisms can mobilize inorganic compounds through autotrophic and heterotrophic leaching,
chelation by microbial metabolites, methylation, and redox transformations (Adeniji 2004).
Microbial alteration of the redox state of either the contaminants or the iron and manganese oxides,
which bind most heavy metals, can make metals and metalloids less soluble. Microbially induced
reduction of hexavalent chromium to trivalent chromium may be the most common application of
bioremediation to metals. Microbial reduction of the highly soluble, oxidized form of selenium to
insoluble elemental selenium by microorganisms is a biological mechanism to remove selenium from
contaminated surface and groundwater (CLU-IN 2008). The adsorption of metals and metalloids onto
microbial biomass can also prevent further migration of these contaminants (Lovley and Coates 1997).
Anaerobic reductive bioremediation to promote bacterial sulfate reduction, and consequent
precipitation of various insoluble metal sulfides, is a possible remediation technique, as demonstrated at
the Stoller Chemical Site in Jericho, South Carolina (CLU-IN 2006).
All of the five anaerobic Superfund projects for metals shown in Figure 3 (and detailed in Appendix A)
treat hexavalent chromium. Two of the five also treat cadmium, manganese, and other metals.
2.3.2 Key Microbe Summary
Bacterial species used in anaerobic reductive bioremediation can be highly specific to a particular
contaminant. For example, Dehalococcoides mcartyii are the only bacteria known to completely convert
PCE to ethene, and some strains also break down polychlorinated biphenyls (PCBs). Dehalobacter spp.
are capable of converting 1,1,1-TCA (known to inhibit the ability of Dehalococcoides to dechlorinate TCEJ
to chloroethane. Pseudomonas stutzeri KC, Methanosarcina barker!, Desulfobacterium autotrophicum,
Moorella thermoacetica, and Methanobacterium thermoautotrophicum are each capable of converting
carbon tetrachloride to methane. Geobacter spp. reduces uranium to a less soluble valence state
(Anderson and others 2003). Perchlorate reducers include the species Dechloromonas aromatic
(Salinero and others 2009), Moorella perchloratireducens (Balk and others 2008), and Sporomusa sp.
(Balk and others 2010). Dissimilatory perchlorate-reducing bacteria (DPRB) are dominated by
Dechloromonas and Azospira spp.
Introduction to In Situ Bioremediation of Groundwater 32
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2.3.3 Sources of Electron Donor
The choice of electron donor (substrate) and the selected delivery method are essential components of
anaerobic reductive bioremediation. Critical considerations include a substrate's properties (solubility,
longevity, cost, and ability to be distributed in the subsurface). Tighter soil matrices can limit the
effective distribution of substrate in the target treatment area. Sometimes a single substrate is not
sufficient, and combinations of substrates may be required. Effective implementation requires careful
design of the mode of delivery and determination of the need for periodic replenishment.
Many materials have been used as electron donors for anaerobic bioremediation. These materials
typically fit into one of two categories: quick or slow release compounds. Table 1 lists some common
substrates and typical applications methods. Many practitioners and commercially available products
use a combination of these types of substrate to capitalize on the advantages of each. In addition,
hydrogen and propane gases are used as electron donors, to a much lesser extent than liquid or solid
substrates.
Table 1. Substrates used for enhanced anaerobic bioremediation (modified from ITRC 2008,
AFCEE 2004)
Soluble substrates
Slow-release substrates
Solid
substrates
(barrier wall
applications)
Substrate
Lactate and butyrate
Methanol and ethanol
Sodium benzoate
Molasses, high- fructose
corn syrup
Whey (soluble)
HRC® or HRC-Xe
Vegetable oils
Vegetable oil emulsions
Mulch and compost
Chitin (solid)
Typical delivery
techniques
Injection wells or
circulation systems
Injection wells or
circulation systems
Injection wells or
circulation systems
Injection wells
Direct injection or
injection wells
Direct injection
Direct injection or
injection wells
Direct injection or
injection wells
Trenching or
excavation
Trenching or
injection of a chitin
slurry
Form of application
Acids or salts diluted in
water
Diluted in water
Dissolved in water
Dissolved in water
Dissolved in water or slurry
Straight injection
Straight oil injection with
water push or high oil/water
content (>20% oil)
emulsions
Low oil content (<10%)
microemulsions suspended
in water
Trenches, excavations, or
surface amendments
Solid or slurry
Frequency of injection
Continuous to monthly
Continuous to monthly
Continuous to monthly
Continuous to monthly
Monthly to annually
Annually to biennially for
HRC (typical), every 3-4
years for HRC-X, potential
for one-time application
One-time application
(typical)
Every 2 to 3 years (typical)
One-time application
(typical)
Annually to biennially,
potential for one-time
application
2.3.3.1 Quick-Release Compounds
Quick-release substrates are typically readily soluble materials and consist of relatively simple molecules
(sugars and alcohols). The solubility nature of the substrate allows for wider distribution into the
subsurface and general movement with groundwater flow. These substrates are rapidly consumed by
Introduction to In Situ Bioremediation of Groundwater
33
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the microbes and require more frequent replenishment. Quick-release compounds include lactate
compounds, organic acids, methanol, ethanol, molasses, and high fructose corn syrup (AFCEE 2004).
2.3.3.2 Slow-Release Compounds
Slow-release substrates have low solubility and higher viscosities than quick-release substrates to make
them generally immobile and sorb to the aquifer matrix. Low solubility and higher viscosities can pose a
challenge for delivery into the subsurface. However, many slow-release substrates are manufactured to
allow products mixed with water to behave as a soluble product during the injection process, most
commonly in an emulsion with surfactants. These materials return to their insoluble nature after a short
period of time (on the order of days) after they are injected. However, breakdown products are soluble
and provide some downgradient distribution of substrate. These substrates consist of long-chain
molecules intended to limit consumption rates and increase longevity of the material in the subsurface.
Longevity is site specific and depends on biological activity, matrix oil retention capacity, and the
hydraulic characteristics of the aquifer. Slow-release substrates include soybean oils (neat and
proprietary emulsified vegetable oil [EVO] products) and proprietary formulations of polylactate esters
and fatty acid esters. Protocol documents are also currently available for design and addition of specific
substrate types, such as edible oil (AFCEE 2007).
Solid substrates, such as mulch, compost, and chitin, are generally the longest-lasting substrates, on the
order of 5 to 10 years (AFCEE 2004). These substrates are best suited for shallow groundwater plumes,
as physical placement of the material is necessary by trenching, excavation, or surface application.
However, chitin can be injected in slurry form for particular applications. Solid substrates are often
replenished by injecting one or more liquid substrates into the solid substrate material matrix. For
example, the mulch biobarriers at the Altus Air Force Base in Oklahoma included piping and other
supporting infrastructure to deliver liquid substrate to the biologically active area to maintain electron
donor concentrations as the mulch decomposed (AFCEE 2008).
Not all anaerobic bacteria can use slow-release substrates. For example, dissimilatory perchlorate-
reducing bacteria (DPRB) cannot break down complex substrates such as edible oils (Coates and Jackson
2008). However, slow-release substrates are gradually consumed by fermenters, which produce simpler
organic compounds like those categorized as quick release compounds. The DPRB are able to consume
these fermentation products, and thus can still be stimulated with either quick- or slow-release
compounds (Borden 2008).
2.3.3.3 Hydrogen and Propane Gas
Direct injection of hydrogen gas is the most direct approach to stimulating anaerobic reductive
bioremediation (AFCEE 2004). Some hydrocarbon gases such as propane can also be used. These
materials are combustible and pose special health and safety and engineering challenges. However,
these gases are generally less expensive than liquid substrates. Delivery methods are similar to those
used to deliver oxygen to promote aerobic bioremediation and, include biosparging and permeable-
membrane diffusers.
Introduction to In Situ Bioremediation of Groundwater 34
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2.3.3.4 Petroleum Hydrocarbons and Other Co-Contaminants
In a mixed contaminant plume, some VOCs such as acetone or petroleum hydrocarbons may be present
in the subsurface as a co-contaminant with oxidized target contaminants (such as halogenated
compounds). In these cases, an electron donor and acceptor are present, but microbial populations able
to use and completely remediate both contaminants may not be present or viable in the aquifer. These
microbial populations will need to be bioaugmented with a consortium of other microbes able to
achieve complete treatment of contaminants. Commonly, co-contaminants will be exhausted before
sufficient remediation of the target compounds has occurred, and additional electron donor may be
necessary.
2.3.4 Delivery Mechanisms
Most liquid substrates are delivered to the subsurface in a solution or mixture with water, either
groundwater or potable water, via injection wells or direct injection using a drilling tool. Applications
anticipated to occur only once may be more cost effective as a direct injection application. However, if
there is any possibility that a site would need more than one application, a permanent injection well
may likely be more cost effective. The well material costs and possible well rehabilitation over the
project life cycle can be less than drilling subcontractor costs, depending on plume depth and time and
pressures needed to inject. The longer the remediation timeframe, the more cost-effective permanent
injection wells become.
High pressure (100 to 500 psi) liquid injections are worth considering for less permeable zones to induce
localized soil fracturing for donor placement. The extent to which soil fracturing will occur is highly
dependent on site conditions. Additional information regarding environmental fracturing is located at:
www.clu-in.org/techfocus/default.focus/sec/Environmental Fracturing/Cat/Guidance and
www.epa.gov/tio/download/citizens/a citizens guide to fracturing for site cleanup.pdf.
Solid substrates are typically applied by excavating into the saturated zone and placing mulch or
compost. These materials are sometimes mixed with a quick- or slow-release substrate and other
amendments. The excavation
is backfilled with earth back to
grade to create a permeable
reactive barrier. However,
chitin can be ground into a
fine powder and mixed with
water for application similar to
liquid substrates. Figure 10
shows an example of a
permeable reactive barrier
configuration.
Figure 10. Permeable Reactive Barrier Example (EPA ZOOlb)
groundwater
being
treated
Introduction to In Situ Bioremediation of Groundwater
35
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2.3.5 Common Byproducts
The most common problematic byproduct of reductive bioremediation is acidity resulting from
fermentation processes. Organic substances are injected to act as electron donors and ensure a highly
reducing environment. These substances are fermented in this environment, which generates hydrogen
ions and, in the absence of adequate buffering capacity, can lower the pH.
Reductive dechlorination produces chloride ions, the reduced form of the chlorine removed from
chlorinated organic compounds. Reductive dechlorination of chlorinated VOCs with multiple
substituents produces intermediate daughter products as each halogen atom is sequentially removed.
An intermediate dechlorination product for a common contaminant such as PCE is VC, which is more
toxic than PCE.
The following lists are some of the most common metabolic byproducts from anaerobic reductive
bioremediation and potential issues associated with those byproducts:
• If nitrate is used, byproducts include nitrite, nitric oxide, nitrous oxide, and nitrogen gas. The
predominant byproduct depends on the enzymes possessed by the microbes present.
• Iron(ll) is far more soluble than iron(lll), so iron reduction could exceed iron water quality
criteria or create a total dissolved solids issue.
• The end product of sulfate reduction is sulfide. If there are not enough dissolved metals to
precipitate the sulfide, hydrogen sulfide gas can be generated, which is toxic.
• Fermentation generates methane, which may necessitate installation of vapor mitigation
systems when a building overlies a treatment area. Organic acids are also generated as part of
fermentation, which can lower the pH of the groundwater and potentially mobilize metals
(notably iron, manganese, and arsenic). The primary concern with mobilization of metals is
creating secondary water quality issues at a site. Monitoring the dissolved metals over time may
be needed to confirm that any mobilized metals precipitate when the pH and ORP return to the
natural state downgradient of the active treatment area. The decrease in pH can also inhibit
Dehalococcoides and stop the bioremediation process altogether.
2.4 Cometabolic Bioremediation
This section contains less detail than sections for other ISB strategies discussed above because limited
full-scale applications of cometabolic bioremediation have been published. Field-scale applications are
planned that will provide information on the utility of this bioremediation strategy. Cometabolism
occurs when microorganisms using one compound as an energy source fortuitously produce an enzyme
that chemically transforms another compound. Organisms thus can degrade a contaminant without
gaining any energy from the reaction. Cometabolic degradation is a process that often happens
concurrently in bioremediation systems designed for direct metabolism of contaminants; however,
some systems have been designed to specifically take advantage of cometabolic processes. Hazen
(2009) indicates that cometabolic bioremediation can occur in environments where contaminant
concentrations are well below concentrations that could provide a carbon or energy benefit to the
biodegrader. Therefore, this method may be effective at degrading very low concentrations of some
contaminants.
Introduction to In Situ Bioremediation of Groundwater 36
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In particular, aerobic microorganisms that degrade methane (methanotrophic bacteria) have been
found to produce enzymes that can initiate oxidation of various carbon compounds. Methanotrophic
bacteria can cometabolize many aliphatic compounds and aromatic compounds (Brigmon 2001).
Cometabolic bioremediation has been shown to degrade contaminants that are typically recalcitrant or
difficult to degrade, such as PCE, TCE, trinitrotoluene (TNT), 1,4-dioxane, and atrazine (Hazen 2009).
Monooxygenase enzymes have shown the ability to oxidize PAHs, PCBs, MTBE, pyrene, creosote, TNT,
NDMA, and 1,4-dioxane (Hazen 2009, Hatzinger 2011, Steffan 2007), and cometabolic bioremediation
has the potential to remediate these groundwater contaminants with further development.
Cometabolic reductive dehalogenation is relevant to large dilute plumes, where contaminant
concentrations are too low for direct reductive dechlorination. Cometabolic reductive dehalogenation
has been observed for PCE, DCA, and CT, and happens concurrently with direct reductive dechlorination,
making it difficult to distinguish the exact contributions of each pathway. Furthermore, there is
laboratory evidence of anaerobic cometabolic degradation of hexachlorocyclohexane, BTEX, PAHs,
atrazine, and TNT, though these remedies have yet to be used extensively in the field. Common
cometabolic bioremediation substrates, enzymes, and contaminants are summarized in Table 2.
Table 2. Common cometabolic bioremediation substrates, enzymes, and contaminants
(from Hazen 2009)
Toluene,
butane,
phenol, citral.
Methane,
Methanol,
Propane,
Propylene
Ammonia,
Nitrate
aldehyde,
cumene, and
limonene Methanol
••••••••••••B
Enzymes
(microbes)
Contaminants
Methane
Monooxygenase,
Methanol
Dehydrogenase,
Alkene
monooxygenase,
catechol
dioxygenase
(Methylosinus)
TCE, DCE, VC, PAHs,
PCBs, MTBE,
creosote, >300
different
compounds
Ammonia
Monooxygenase
{Nitrosomonas,
Nitrobacter)
TCE, DCE, VC,
TNT
Toluene
Monooxygenase,
Toluene
Dioxygenase
(Rhodococcus,
Pseudomonas,
Arthrobacter)
TCE, DCE, VC,
1,1-DCE, 1,1,1-
TCA, MTBE
Alcohol
Dehydrogenases
(Pseudomonas,
Strep tomyces,
Corynebacterium)
PCE, TCE, DCE,
VC, Hexachloro-
cyclohexane
Dehalogenase,
AtzA,
Dichloromethane
Dehalogenase
(Dehalococcoides,
Methanogens,
Desulfovibrio,
Clostridium,
Geobacter,
Ciavibacter)
BTEX, PCE, PAHs,
Pyrene, Atrazine,
TNT, etc.
Introduction to In Situ Bioremediation of Groundwater
37
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3.0
Treatability studies are commonly performed during or after development of remedial alternatives
(feasibility study) where bioremediation is a potential site remedy. Treatability studies generally include
bench-scale or pilot-scale studies and are used to further evaluate whether the proposed
bioremediation remedy will be successful under site conditions. These studies also provide the design
information required for full-scale implementation. Once designed and installed, a bioremediation
system will require process and performance monitoring and possible modification to optimize the
bioremediation system. The following sections provide design considerations for bench testing, pilot
tests, and full-scale implementation. Delivery and Mixing in the Subsurface: Processes and Design
Principles for In Situ Remediation (Kitanidis and McCarty 2012) provides a more detailed discussion of
many topics in this section.
3.1 Treatability Studies
Bench-scale studies or pilot tests are important pre-design steps in determining whether a remedial
technology is an implementable, cost-effective, and scalable treatment option. Successful bench-scale
studies or pilot tests can justify implementing a full-scale bioremediation treatment system, while failure
of bench or pilot tests may indicate that the bioremediation design needs to be reconsidered or possibly
abandoned. Treatability studies are not unique to bioremediation, and the general approach to
completing these studies is consistent among most in situ remedial technologies. There are several
reference documents that present specific treatability study methods for various bioremediation
strategies (AFCEE 2004, Stroo 2009).
3.1.1 Bench Test
Bench testing is performed on a sample of the site soil, groundwater, and bedrock, if present, collected
for use in laboratory-scale treatment studies. Bench tests are generally used to evaluate:
• The performance of various amendments (biostimulation);
• Substrate demand and loading rates that can subsequently be tested in a pilot test;
• Whether addition of bacterial culture is needed (bioaugmentation);
• What consortium or combination of bacterial cultures is optimal; and
• The treatability of contaminants at different concentrations.
Bench testing is not typically performed for aerobic bioremediation studies and is often referred to as a
microcosm study for anaerobic projects. Bench testing allows for easier manipulation and testing of
many variables. If site conditions are favorable for a particular bioremediation approach, the cost of a
bench test may outweigh the benefits. However, if site conditions are marginal, a bench test could be
useful in evaluating whether bioremediation can be applied at a site before additional investment is
made in pilot testing (AFCEE 2004). Bench testing may not always accurately reflect subsurface
conditions in the field; however, downhole forms of microcosms, called biotraps, can also be used in the
field as an alternative to bench studies. A biotrap is a passive sampling tool containing a matrix that
encourages colonization by subsurface microbes. Biotraps can be used to test different amendments
and microbe consortiums on a microcosm scale and calculate degradation rates.
Introduction to In Situ Bioremediation of Groundwater 38
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3.1.2 Pilot Test
Pilot tests are usually small-scale field tests and typically include a set of injection wells or direct
injection points and monitoring wells at varying distances within the pilot test treatment area.
Monitoring wells may be positioned radially around the injection area when groundwater flow velocities
are low, or at various downgradient distances when groundwater flow velocities are naturally higher or
a circulation cell will be established. When possible, a tracer test using a conservative tracer, such as an
ion salt (such as sodium bromide) or a dye (for example, fluorescein or rhodamine), should be
completed as part of the start of the pilot test to help determine groundwater flow paths, dispersion,
effective porosity, and velocity.
The results from pilot tests will help identify microbial response to biostimulation and provide design
data regarding radii of influence of the injection wells and the performance of the amendment. Pilot test
results are used to establish the full-scale injection well spacing and depth interval, as well as the
quantity of amendment and the frequency of application. Pilot tests can also be used to evaluate
bioaugmentation. A pilot test is generally designed with scalability in mind. For example, the designer
will consider whether a given pilot test layout could be scaled up to treat the total target area within the
cleanup and cost parameters of the project. In addition, a designer will decide whether other factors
need to be considered, such as source water and electricity demands for pumps that may be in more
remote locations during full-scale implementation. In some cases, a bench test will not be required if a
pilot test is well designed (AFCEE 2004).
If bench testing is not completed, a pilot test may examine the effectiveness of multiple amendments
and nutrients (biostimulation) and bacterial cultures (bioaugmentation), rather than simply confirming
the results of the bench test. However, pilot testing of multiple amendments or cultures requires some
design considerations to differentiate results. One design would be to test various amendments or
cultures in different portions of the pilot test area separated by an unamended control area. Another
design would be to test amendments sequentially. For example, Kovacich and others (2006)
implemented a pilot test to evaluate anaerobic reductive dechlorination of a plume of TCE in
groundwater. The pilot test was designed to deliver sodium lactate to a circulation system. Electron
donor delivery problems resulted in poor performance and incomplete dechlorination. A second phase
of the pilot test was implemented to evaluate direct injection of EVO to the area of the circulation
system between the injection and extraction wells. The second phase of the pilot test ultimately showed
that the technology could be applied to the site.
Pilot tests should be conducted for a period long enough to determine if complete biological
degradation is achieved in addition to obtaining the design data previously mentioned. However, longer
pilot tests could provide important information regarding amendment longevity, longer-term microbial
and aquifer geochemical responses, potential maintenance issues, and contaminant rebound
characteristics.
Introduction to In Situ Bioremediation of Groundwater 39
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3.1.2.1 Biostimulation
Amendments can be applied to sites in
various ways. Figure 11 shows three
common application methods.
Biological groundwater amendments
have been applied using direct-push
injection tooling, permanent injection
wells (vertical and horizontal),
infiltration trenches, and permeable
reactive barriers (PRBs).
Determination of amendment
quantities is highly dependent on the
specific biological process targeted to
achieve remediation. Required
quantities can be calculated based on
stoichiometry, estimates of biological
demand, or rules of thumb found in
guidance documents, literature, or
provided by product vendors. Another
consideration in calculating amendment
quantities is the application rate.
Application rates can be based on
results of bench testing various
contaminant concentrations. In the
absence of bench testing, application of
amendments at a uniform rate across a
target treatment area still may not be
the best approach. A more cost-
effective methodology is often to focus
the amendment in the zones
(horizontally and vertically) of highest
concentrations to address the areas of
greatest flux and to reduce the
application rate on the margins of the
plume. It is important to recognize that
the highest concentrations may not be
in the most transmissive zones. Full-
Direct
Injection
PRB
Circulation
Figure 11. In situ Bioremediation System
Configurations (EPA 2000), as adapted
scale design considerations for source area or dissolved plume treatment are discussed in Section 3.2.
Various amendment quantities can be identified for different zones and injection infrastructure can be
designed for each zone. High-resolution vertical profiling of the aquifer characteristics and contaminant
distribution would result in a more precise and targeted design.
Introduction to In Situ Bioremediation of Groundwater
40
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3.1.2.2 Bioaugmentation
A site is typically bioaugmented after an aquifer has been biostimulated and favorable conditions for the
target microbial community exist. MBTs can be utilized before, during, and after biostimulation to
determine if bioaugmentation is necessary. If a site is selected for bioaugmentation, then
bioaugmentation is usually tested as part of a pilot test. During the pilot test, inoculation with multiple
cultures (in unique portions of the pilot test area) can be performed to evaluate each culture's
performance and confirm bench testing with field data. MBTs can be used throughout the pilot-testing
phase of a project to assist in evaluating whether the bioremediation application is performing as
designed, how the microbial community changes over time given the electron donor and acceptor use
and contaminant degradation, and whether the technology will likely achieve the remedial objectives
(ITRC2013).
3.2 Full-scale Implementation
Results from bench and pilot tests are used to guide full-scale design and implementation. A full-scale
implementation approach is selected considering the site CSM, remedial goals, regulatory requirements,
and future site use or development to design a remedial system that best meets the needs of the site
with the lowest total life cycle cost. Design tools are available to assist with full-scale implementation.
For example, several design tools were developed as part of project ESTCP ER-200626.
Figure 12 presents bioremediation treatment approaches applied at NPL sites from 1989 to 2008. For
amendment delivery, 41 of the operating or completed projects listed in Appendix A use or used direct
Unknown orTBD
1
1%
Recirculation
13
15%
Biobarrier & Recirculation
3
3%
Biobarrier
23
27%
Note: Includes remedies selected in
FY 1989-2008decision documents.
Refer to Appendix for data.
TBD = To be determined.
Direct Injection
26
_ Direct Injection &
Biobarrier
4
SK
Direct Injection &
Recirculation
11
13%
Figure 12. Bioremediation Design Types
Introduction to In Situ Bioremediation of Groundwater
41
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injection or biobarriers/PRBs and 27 use or used groundwater recirculation. These categories are not
mutually exclusive, and a single project may be counted in more than one category. Bioaugmentation
was used at 18 of the projects in Appendix A. This section addresses three primary approaches to full-
scale implementation that include active, semi-passive, and passive (Stroo 2009).
3.2.1 Active Treatment Approach
Active treatment approaches to bioremediation include circulation of groundwater in the target
treatment area, as shown in Figure 13. Circulation requires significant capital cost to install the
extraction wells, injection wells, associated conveyance lines, and the system building or skid where
amendments will be stored and metered into the delivery system. Frequent operation and maintenance
(O&M) is required to make system checks and adjustments. Circulation also requires a continuous
source of power to run the circulation pumps. Water-soluble amendments or amendments that are
emulsified are required for effective distribution throughout the target treatment area of a circulation
system. Active systems can also distribute bacterial culture, if bioaugmented, much more effectively
than semi-passive or passive methods. Circulation approaches can effectively treat target areas in less
time and, as a result, may have lower total life cycle costs. Paired injection and extraction typically
increases the hydraulic gradient at a site, thus increasing the rate of distribution and delivery of
amendments. Active treatment is often applied to source areas and highly concentrated, smaller
GROUNDWATER FLOW PATH
INJECTION WELL PW,eA MONITORING WELL
® EXTRACTION WELL
EW-1
Figure 13. Typical Circulation System Layout (Kovacich and others 2006)
Introduction to In Situ Bioremediation of Groundwater 42
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dissolved plumes where elimination of significant mass in a short time can provide the best value to an
overall treatment program.
3.2.2 Semi-Passive Treatment Approach
Semi-passive treatment approaches to bioremediation are similar to active approaches. They also
include circulation of groundwater and require the infrastructure and water soluble donors listed above,
but are typically applied at sites where longer treatment times are acceptable. The primary difference
between semi-passive and active systems is that semi-passive systems are not operated continuously.
Amendments are circulated throughout the target treatment area in pulses. For example, amendment
might be circulated for 3 months, perhaps long enough to circulate one pore volume within the target
treatment area. Circulation is halted, and the site is monitored for a time to determine when additional
amendments are required. Semi-passive treatment may take more time than active treatment, but may
result in less energy consumption, less O&M, and less use of amendment. As a result, semi-passive
treatment approaches may have a lower total life cycle cost than active approaches.
3.2.3 Passive Treatment Approach
Passive treatment approaches to bioremediation differ from active and semi-passive approaches in
several ways. Groundwater is not circulated for extended periods, slow-release amendments are more
often used, relatively little infrastructure is required, and treatment times can be longer. However,
amendments and infrastructure can vary widely for passive approaches based on aquifer properties and
amendment cost and longevity. Passive treatment approaches rely on natural flow of groundwater to
deliver contaminated groundwater to biologically active areas where treatment occurs. The three most
common treatment area amendment delivery configurations include a grid of injection points, a line of
injection points, or a trench filled with substrate (a PRB), as shown in Figure 14. These treatment
configurations and designs have been detailed in recent protocol documents from the Air Force Center
for Engineering and the Environment (AFCEE, 2004, 2007, and 2008) and the Interstate Technology and
Source Area
Containment
Barrier (Biobarrier)
©
Groundwater
Flow
Permeable Reactive Barrier
Injection Point
Figure 14. Schematic of Source Area and Barrier Injection Configurations.
(Adapted from AFCEE 2004)
Regulatory Council (ITRC 2005). These configurations can also be used in combination with one another
Introduction to In Situ Bioremediation of Groundwater
43
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and even in conjunction with semi-passive and active approaches. Additional details regarding each
configuration are presented in the following sections.
3.2.3.1 Treatment Area Grid Configuration
Treatment area grids are used to address source areas and smaller dissolved plumes in groundwater, as
shown in Figure 14. Treatment areas are often completed by direct-push injections in a grid pattern or
by establishing a temporary circulation system to distribute substrate and bacteria, sometimes referred
to as biozones. Generally, more closely spaced wells will increase drilling costs but reduce the duration
of an injection event. Conversely, wells spaced farther apart will decrease drilling costs but increase the
duration of an injection event. For example, the volume of amendment and water required to achieve
coverage for each well in a grid spacing of 20 feet is four times the volume required for a grid spacing of
10 feet. Assuming the same injection rate, the time required would increase by a factor of four. (So, for
example, an injection into wells with a 10-foot spacing that takes 6 hours would take 24 hours using
wells with a 20-foot spacing.) Selecting the most cost-effective grid spacing for each site is highly
dependent on drilling and amendment delivery implementation costs and is evaluated on a case-by-case
basis. The primary cost drivers can be analyzed to identify the most cost-effective spacing, as illustrated
in Figure 15, where the optimal well spacing is 12.5 feet.
Barrier Cost vs Well Spacing
5160,000
5140,000
o
o
.Well Installation
Costs
-Injection Costs
Substrate Costs
• Total Installation
and Injection Costs
10 15 20
Well Spacing (feet)
Figure 15. Example Cost Comparison for a PRB with Various Injection Well Spacings
(ESTCP 2006)
3.2.3.2 Migrating Plume Barrier Configuration
Passive treatment zones that are created perpendicular to the axis of the plume are often called
biobarriers or PRBs. Biobarriers are often applied at sites with large dissolved plumes where active and
semi-active approaches are not cost effective. Biobarriers are also used to provide treatment and
containment of the plume to prevent off-site migration or discharge to a vulnerable receptor.
Introduction to In Situ Bioremediation of Groundwater
44
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3.2.3.3 Circulation Configuration
Circulation systems are one of the most efficient ways to distribute amendments and bacteria and are
typically used in active and semi-passive treatment approaches. In a passive treatment approach,
circulation systems can be used to distribute emulsified slow-release amendments. Some slow-release
amendments remain suspended in emulsion for 5 to 10 days after they are injected. Circulation allows
for greater hydraulic control and manipulation (in other words, induces a greater hydraulic gradient) and
can be used to create treatment areas beneath buildings, active roads, runways, and other areas with
limited site access. However, the total area that can be circulated is limited by the time the amendment
remains suspended.
3.2.4 Vertical Applicatio n and Distributio n
Implementation costs related to almost any technology increase with greater depth and treatment
thickness. When injection wells are used, initial injection rates are typically similar to the theoretical
transmittance capacity of the screens and are adjusted as needed in the field after startup to ensure
distribution across the entire screened interval. Careful consideration for injection well screens is
important. Large-diameter wells with high-flow screens may seem to be the best approach to inject
fluids, but this approach may not result in effective distribution of amendments. For example, if a
selected screen has a theoretical transmittance capacity of 2 gallons per minute (gpm) per foot and its
total length is 10 feet, the overall theoretical transmittance capacity is 20 gpm. If the designed injection
rate on that screen is only 5 gpm, there is a possibility that the screen will not be fully pressurized and
injection materials will enter through the top 2.5 feet of screen or the most permeable interval. Well
diameters and screen characteristics (openings size and type [slot versus continuous wrap]) need to be
specifically designed with the aquifer material and target injection rate in mind. Once installed,
thorough well development is required to maximize injection efficiency.
In a single injection point, multiple screens of shorter lengths (perhaps 5 to 10 feet long) may be
required to achieve adequate vertical distribution, rather than one long screen (for example, 20 to 50
feet long). Injection wells intended to provide multiple injection depths are commonly installed as well
clusters, sometimes bundeled in the same borehole or as individual wells in separate boreholes but
closely spaced together. Figure 16 shows an example of an injection well design with multiple screened
intervals within a single borehole that is currently in use at a passive bioremediaton site. The optimal
injection well design will be site specific. The well cluster approach may have higher associated drilling
costs but result in lower injection costs because downhole packers or other equipment are not needed.
Introduction to In Situ Bioremediation of Groundwater 45
-------
5
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NOTE:
A DOWN-HOLE K-PACKER WILL BE UTILIZED TO ISOLATE
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Figure 16. Example Injection Well Design. (Courtesy of Tetra Tech, Inc.)
3.2.5 Maintenance
Re-application of amendments, including electron acceptor or donor, will be required at most sites.
Some sites may require geochemical adjustment and nutrient amendments. Maintenance applications
are sometimes overlooked during the initial design, and consideration of future drilling costs associated
with maintenance applications may require installation of permanent wells. New wells with shorter or
focused screens or shorter target injection intervals may be needed in areas where data suggest
amendment distribution is inadequate.
The success of biological technologies depends on the presence and persistence of the amendment
(electron acceptor or electron donor) and maintaining the geochemical conditions in groundwater that
will allow biological populations to flourish. As discussed in Section 1.5.4.2, pH in an aquifer is critical to
the performance of bioremediation systems. Adjustment of pH during subsequent maintenance
applications, typically with the addition of a base like sodium bicarbonate or sodium hydroxide may be
needed; however, pre-design data should inform a practitioner of the potential need for post-
installation pH adjustment. An effective performance monitoring program, as discussed in Section 3.3, is
required to decide when additional amendments are required to maintain the biologically active zone.
Introduction to In Situ Bioremediation of Groundwater
46
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Before amendments are applied to a site, the injection wells must be evaluated to assure they can still
operate at their designed injection rates.
Well screen fouling can be a major maintenance issue for aerobic treatment, including calcium and iron
precipitation and biological fouling by iron bacteria. Injection well maintenance typically includes well
development to remove precipitates and biological films on the screen and gravel pack and usually
involves chemical treatment with acids. Biomass buildup in injection wells can also be problematic to
reductive dechlorination sites if too much substrate is applied and injection wells are inadequately
flushed with water after injection. General maintenance may include use of a downwell video camera to
monitor fouling, periodic cleaning with weak organic or inorganic acids, biocides, bleach, or chlorine
dioxide for biomass, and well re-development.
3.3 Measuring Performance
As with any site cleanup, it is important to measure progress toward the remedial objectives1.
Measuring remedial performance is critical to its optimization and long-term applicability to the site.
Depending on the remedial objectives, it may take a few years or decades to reach remedial objectives.
Key questions to ask when a monitoring approach is developed include (ITRC 2011):
• What media should be monitored?
• What constituents should be monitored?
o Beyond the contaminants of concern (COCs), what other parameters should be
monitored to establish multiple lines of evidence to evaluate performance?
o How many lines of evidence are needed for an assessment toward an objective?
• What metrics should be used?
• Where should monitoring points be located?
• When should monitoring occur?
With biologically dependent remedies, the initial monitoring is critical to the overall success of the
remedy to ensure aquifer geochemistry has responded as bench and pilot testing suggested; target
bacterial communities are established, sustained, and thriving; and to track the general biogeochemical
responses to the initial installation. For example, initial performance monitoring could be monthly for a
quarter, followed by quarterly for a year, and semi-annually thereafter. It is important to evaluate
performance data after each injection event to ensure that expected trends are observed. Even at sites
where initial performance is promising, the groundwater monitoring program must consider the
possibility of rebound caused by various factors, including matrix diffusion (as described in
Section 1.6.3.4).
1ForSuperfund program information on monitoring performance and progress of groundwater remedies, see
www.epa.gov/superfund/health/conmedia/gwdocs/pdfs/gwroadmapfinal.pdf
Introduction to In Situ Bioremediation of Groundwater 47
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The expected trends vary for each bioremediation strategy, but generally include reductions in
contaminant concentrations, stable redox conditions, stable geochemistry, and adequate amendment
concentration. However, a short-term increase in contaminant mass may occur initially after installation
as a result of changes in the equilibrium between the contaminant phases (adsorbed, dissolved, and
NAPL) and liberation of contaminant into the dissolved phase. The following conditions may indicate
poor performance of a bioremediation remedy:
• Limited, incomplete, or no reduction, and even increases in contaminant concentrations
(although temporary increases in contaminant concentrations can be expected near source
zones as a result of increased dissolution);
• Trending toward aerobic conditions in an anaerobic remedy and trending toward anaerobic
conditions in an aerobic remedy;
• Increasing concentrations of competing electron acceptors or donors, for anaerobic and aerobic
remedies; and
• Sharply decreasing concentrations of amendments or amendment concentrations below those
necessary to support bioremediation at a site.
When these conditions occur, modifications in the system may be required to improve performance.
Possible modifications may include changes in the method of amendment delivery, changes in the
selected amendment, and conditioning the aquifer geochemistry.
Performance monitoring at aerobic bioremediation sites typically tracks three key indicators:
• The concentration of oxygen and compounds being used as a source of oxygen. Dissolved
oxygen can be measured in the field;
• The redox conditions within the aquifer and concentrations of primary terminal electron
acceptors. ORP measurements in the field can provide the general redox state of an aquifer;
and
• The concentration of contaminants and daughter products.
Microbial testing is not commonly conducted with most aerobic bioremediation sites, which are
associated with hydrocarbons.
Performance monitoring at anaerobic bioremediation sites typically tracks four key indicators:
• The concentration of competing electron acceptors and resulting reduced states as an indicator
of redox conditions. ORP measurements in the field can provide the general redox state of an
aquifer;
• The concentration of organic carbon (total and dissolved) and substrate breakdown products,
such as volatile fatty acids, to make sure adequate donor is present and in a useable form for
the target bacteria populations;
• The concentration of contaminants and dechlorination daughter products. Evaluation of the
basis of molar concentration provides insight on the conversion of contaminant mass. Various
dissolved hydrocarbon gases, such as ethene and ethane, require specialty analysis other than
Introduction to In Situ Bioremediation of Groundwater 48
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the standard VOC scan. Methane is also included in that analysis, which provides further
information on the redox conditions; and
• The concentration of the target bacterial populations. Groundwater or filter media through
which groundwater has passed can be processed for deoxyribonucleic acid (DNA) sequencing by
qPCR or possibly other MBTs (ITRC 2013). qPCR can quantify target bacterial groups (such as
Dehalococcoides), as well as quantification of a subset of Dehalococcoides with genes that can
perform vinyl chloride reduction to ethene, vinyl chloride reductase (vcrA).
A performance monitoring program is intended to measure performance at key locations within a
plume. Generally, wells would be located within the biologically active area, immediately downgradient
of the biologically active area, and farther downgradient at distances based on site seepage velocities,
monitoring frequency, and any regulatory requirements.
Overall, densities of targeted bacterial populations should increase with time and reach optimal levels
(>107 cells/L), geochemical conditions must remain favorable, and contaminant levels should decrease in
all performance monitoring wells and eventually in downgradient point of compliance wells for
successful bioremediation applications. If progress toward remedial objectives is not adequate,
reevaluation of a remedy could be warranted (ITRC 2011, EPA 2011).
Introduction to In Situ Bioremediation of Groundwater 49
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4.0
The field of bioremediation is still a relatively young discipline, and new developments occur each year.
The following sections provide a summary of some emerging trends.
4.1 Environmental Remedy Footprint
Increasing attention is being paid to the environmental footprint of implementing a remedy. An
environmental footprint includes energy usage, air emissions, water usage, materials usage, and waste
generation. ISB treatment options often result in a smaller environmental footprint onsite than ex situ
or non-biological methods. For example, ex situ options are energy intensive because of the need to
remove and transport large quantities of soil or groundwater. ISB, however, allows treatment without
transportation of the contaminated media. The approach to implementing an ISB remedy also can have
significant effects on the environmental footprint of the remedy. For example, using extracted
groundwater to blend and inject the electron donor has a lower water footprint than using potable
water for this purpose. Furthermore, using an electron donor that is a food-grade byproduct or waste
product from the food preparation industry can have a lower footprint than using a specially prepared
electron donor. Using multiple, long-term direct-push events may have a larger energy and air emission
footprint than using permanent injection wells. Thorough consideration of the appropriate design
parameters for a successful remedy and consideration of the remedy components that contribute the
most to the remedy's environmental footprint can lead to a reduction in the footprint and successful
remedies (EPA 2012b).
4.2 Compound Specific Isotope Analysis
The following section references the EPA document, "A Guide for Assessing Biodegradation and Source
Identification of Organic Ground Water Contaminants using Compound Specific Isotope Analysis (CSIA)."
CSIA is an environmental forensics technique used to characterize contaminated sites and track the
progress of bioremediation and natural attenuation. CSIA measures and compares the ratios of stable
isotopes found in compounds of suspected contaminant sources or plumes as well as the feedstock or
manufacturing process of materials historically used in the vicinity of the site. Isotopic analysis can help
identify various sources of the same compound based on their different isotopic "signatures." It also can
be used to evaluate the extent of contaminant degradation caused by microbes. Typical forensic stable
isotopes include carbon, hydrogen, chloride, sulfur, and oxygen. However, the majority of the work is
done with carbon isotopes (EPA 2008).
There are several techniques to study biodegradation in groundwater that involve the addition of
contaminants that are artificially labeled with a carbon isotope (usually 13C-label). Examples include
stable isotope probing (SIP) and Bio-Sep beads (media) amended with ISC-labeled substrates. The
media can be placed in groundwater wells to conduct in situ SIP studies. The isotopically-enriched
contaminant is applied to the medium and the medium is incubated in a well for a given period of time.
These techniques work in much the same way as radiocarbon labeling; the 13C-label is used to track the
transfer of carbon from the substrate to its metabolites, or to the dissolved inorganic carbon pool, and
its subsequent incorporation into the microbial biomass. The disappearance of the label from the
substrate pool is convincing evidence that the targeted compound is indeed degrading, and the
Introduction to In Situ Bioremediation of Groundwater 50
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identification of 13C-label in microbial biomass is definitive proof that the compound was biologically
degraded (EPA 2008). The ITRC Environmental Molecular Diagnostics Fact Sheets released in April 2013
(ITRC 2013) also discuss compound-specific isotope analysis and SIP in detail and provide examples of
applications and limitations.
4.3 High-Resolution Site Characterization
Effective implementation of remedial technologies, especially in situ methods such as ISB, requires
accurate site characterization. In particular, the use of HRSC can vastly improve the CSM. HRSC has
become more prominent as sampling techniques, data evaluation, and presentation methods have
improved. HRSC strategies and techniques use scale-appropriate measurement and sample density to
define contaminant distributions, and the physical context in which they reside, with greater certainty,
supporting faster and more effective site cleanup (CLU-IN 2013). The data obtained from HRSC are used
to develop an accurate CSM by identifying heterogeneities in the subsurface that significantly influence
contaminant distribution, fate, and transport. These heterogeneities can occur at very small scale that
conventional investigation strategies and technologies (primarily placing monitoring wells at biased
locations to delineate extent of contamination) can miss.
HRSC uses transects of vertical subsurface profiles oriented perpendicular to the direction of
groundwater flow. Profiles located along each transect are used to collect high-resolution lithologic,
hydrogeologic and contaminant data using real-time direct sensing tools implemented using direct push
technology (DPT). Lithologic data are collected using such technologies as cone penetrometer testing
(CPT), various electrical conductivity (EC) probes, and hydraulic profiling tools. The hydrogeologic data
are best provided by real-time hydraulic profiling tools. Contaminant data are provided using such
technologies as Laser Induced Fluorescence (LIF), Membrane Interface Probe (MIP), and Tar-specific
Green Optical Screening Tool (TarGOST®).
4.4 3-D Visualization and Analysis of Data and In Situ Sensors
Several software programs are available to perform 3DVA of site characterization and performance
monitoring data. These programs are useful for designing amendment delivery systems and identifying
which portions of a plume may require additional amendments. Some practitioners are combining in
situ sensors (pH, dissolved oxygen, ORP, chloride, and conductivity) and a web-based interface to
facilitate continuous monitoring and evaluation additional amendment needs.
The 3DVA programs typically use geostatistical kriging procedures to establish the spatially-accurate
distribution of each parameter in three dimensional space. Figure 17 presents an example of three-
dimensional kriging using CTech Corporation's Mine Visualization System (MVS) software. Integrated
visualizations can be made when HRSC and performance monitoring data are combined, providing
increased understanding of contaminant distribution and behavior. Visualizations can be fully
articulated to enable site conditions to be viewed from any vantage point of interest, allowing for more
rigorous analysis. However, the visualizations are only as accurate as the data used to prepare them and
the software skill and geostatistical knowledge of the modeler. Uncertainty will always exist between
data points. Characterizing sites using HRSC will increase data density and reduce uncertainty.
Introduction to In Situ Bioremediation of Groundwater 51
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Figure 17. Example of Geostatistical Kriging Analysis of Multi-depth TCE Soil Concentrations.
(Courtesy of Tetra Tech, Inc.)
Introduction to In Situ Bioremediation of Groundwater
52
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5,0
ISB can be applied to the treatment of source materials or plumes, and has proven applicable to
numerous inorganic and organic contaminants. Several different biochemical pathways have been used,
numerous amendments have been developed, and numerous methods have been implemented to
deliver those amendments. Implementation of ISB is highly flexible, often using one or a combination of
active, semi-passive, or passive delivery systems. Amendments can be liquid, solid, or gaseous to serve
as electron donors, acceptors, cometabolites, and nutrients. Table 3 provides a summary of ISB
strategies presented in this document. Clear guidance documents on the most widely used forms of ISB
have been developed and are widely available. However, site-specific testing is usually appropriate
before the final design is completed to help identify the optimal amendment type, amendment quantity
and delivery system.
Effective implementation of ISB often requires careful monitoring with the potential for adjustments to
the amendments and the delivery system. After treatment, several years may be required before
conditions re-equilibrate to pre-treatment and pre-impact conditions. Finally, innovation continues and
several emerging trends will affect the selection, design, and operation of ISB systems in the future.
Introduction to In Situ Bioremediation of Groundwater 53
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Table 3. Summary of ISB strategies
ISB Strategy
Key
Characteristics
Target
Contaminants
Advantages
Limitations
Aerobic
Relies on
presence of
oxygen
Petroleum
hydrocarbons
and some fuel
oxygenates
Ionic form of
metals
Widespread
acceptance with
documented
success for
treating target
contaminants
Aerobic bacteria
responsible for
degradation are
generally
ubiquitous in
nature
Some petroleum
derived plumes
are very
reduced
requiring high
doses of oxygen
Delivery systems
may encounter
significant
biological
fouling
Anaerobic
Oxidative
Relies on
addition or use
of other
electron
acceptors
besides oxygen
Petroleum
hydrocarbons
present in
reducing
conditions
May be applied
to highly
reduced
plumes
Limited use to
date
Can be difficult
to distinguish
from
microaerophilic
oxidation
Anaerobic
Relies on electron
donor additions
uses contaminants as
electron acceptors
Anaerobic
metabolism includes
fermentation,
methanogenesis,
reductive
dechlorination,
sulfate- and iron-
reducing activities,
and denitrification
Chloroethenesand
chloroethanes
Perchlorate,
Munitions,
Chromate, and
Nitrate
Widespread
acceptance with
documented success
for treating target
contaminants
Documented success
in high concentration
source material
Abiotic degradation
often occurs parallel
to biological
degradation
processes
Sensitivity to specific
range of geochemical
conditions
May require
bioaugmentation
with commercially
available microbial
cultures
Aerobic
Cometabolism
Relies on addition
of cosubstrates for
fortuitous
degradation of
contaminants
May be used under
aerobic or
anaerobic, based on
the redox state of
the contaminant
May be applicable
to:
PAHs, Explosives,
Dioxane, NDMA,
PCBs, Pesticides,
MTBE,
Chloroethenes,
Chloroethanes,
Chloroform, and
methylene chloride
Maybe able to
treat contaminants
to low cleanup
levels
Limited use to date
in field applications
Inhibitory
intermediate
products can be
produced
Substrate pulsing
may be needed to
reduce competitive
inhibition between
use of substrate
and contaminant by
the microorganisms
Introduction to In Situ Bioremediation of Groundwater
54
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Bradley, P.M. 2007. Dichloroethene and vinyl chloride degradation potential in wetland sediments at
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Brigmon, R.L. 2001. Methanotrophic Bacteria: Use in Bioremediation. Prepared for the U.S. Department
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Brubaker, G.R. and H.F. Stroo. 1992. In situ bioremediation of aquifers containing polyaromatic
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CL:AIRE SABRE Bulletin SAB 1. 2010. Project SABRE (Source Area BioRemediation)-an Overview.
CLU-IN. 2006. Technology News and Trends. February. On-line address:
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LINKS TO ADDITIONAL INFORMATION
EPA. Use of Bioremediation of Superfund Sites (542-R-1-019):
www.epa.gov/tio/download/remed/542r01019.pdf
EPA Superfund Remedy Report, 13th edition:
www.clu-in.org/download/remed/asr/13/SRR 13th MainDocument.pdf
www.clu-in.org/download/remed/asr/13/SRR 13th Appendices.pdf
EPA Superfund Remedy Report, 12th edition:
www.clu-in.org/download/remed/asr/12/asrl2 main body.pdf
www.clu-in.org/download/remed/asr/12/asrl2 print appendices.pdf
www.clu-in.org/download/remed/asr/12/asrl2 online appendices.pdf
EPA. Monitored Natural Attenuation of Inorganic Contaminants in Ground Water, Volume 2:
http://nepis.epa.gov/Adobe/PDF/60000N76.pdf
Introduction to In Situ Bioremediation of Groundwater 62
-------
Federal Remediation Technologies Roundtable (FRTR). Remediation Technologies Screening Matrix and
Reference Guide, Version 4.0:
www.frtr.gov
ITRC Bioremediation of DNAPLs Documents:
www.itrcweb.org/guidancedocument.asp?TID=47
ITRC Enhanced In situ Biodenitrification Documents:
www.itrcweb.org/guidancedocument.asp?TID=74
ITRC In situ Bioremediation Documents:
www.itrcweb.org/guidancedocument.asp?TID=9
EPACIu-in, Bioremediation
www.clu-in.org/techfocus/default.focus/sec/Bioremediation/cat/Overview/
EPACIu-in, Bioremediation of Chlorinated Solvents:
www.clu-in.org/techfocus/default.focus/sec/Bioremediation of Chlorinated Solvents/
cat/Overview/
EPACIu-in, Bioventing and Biosparging:
www.clu-in.org/techfocus/default.focus/sec/Bioventing%5Fand%5FBiosparging/cat/Overview/
SERDP/ESTCP, Bioremediation Documents:
www.serdp-estcp.org/Program-Areas/Environmental-Restoration
For additional information on this document, please contact:
Linda Fiedler or Edward Gilbert
fiedler.linda@epa.gov gilbert.edward@epa.gov
Introduction to In Situ Bioremediation of Groundwater 63
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Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
^^•_
EASTLAND WOOLEN
MILL
FORT DEVENS - OU8 -
AOC50/PCE Spill
HANSCOM
FIELD/HANSCOM AIR
FORCE BASE - Site 1
On-Site Plume
INDUSTRI-PLEX-
Groundwater
impacted by West
Hide Pile
PARKER SANITARY
LANDFILL - OU1
UNION CHEMICAL
CO., INC. -OU 1
BOG CREEK FARM
BRIDGEPORT RENTAL
& OIL SERVICES -
Deep Groundwater
CHEMICAL CONTROL -
In situ Bio
COLESVILLE
MUNICIPAL LANDFILL
- In situ
Bioremediation
EMMELL'S SEPTIC
LANDFILL
FEDERAL AVIATION
ADMINISTRATION
TECHNICAL CENTER
(USDOT) - OU 1, Area
D -Jet Fuel Farm-
Near MW-19S
OU
01
08
01
02
01
01
02
02
00
01
02
01
Document
Type
ROD-A
ROD
ROD
ROD-A
ROD
FYR
ESD
ROD
EPA ID
MED980915474
MA7210025154
MA8570024424
MAD076580950
VTD981062441
MED042143883
NJD063157150
NJD053292652
NJD000607481
NYD980768691
NJD980772727
NJ9690510020
EPA
Region
01
01
01
01
01
01
02
02
02
02
02
02
Bioremediation
Type
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Biosparging
In situ Aerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
N/A
Biobarrier
Direct injection
N/A
Biobarrier
Unknown orTBD
N/A
N/A
Direct injection
Biobarrier
N/A
Recirculation
Contaminants
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
BTEX; Nonhalogenated
VOCs
BTEX; Halogenated
VOCs; Nonhalogenated
VOCs
Halogenated VOCs
Halogenated VOCs
BTEX; Halogenated
VOCs; Nonhalogenated
VOCs
Halogenated VOCs;
Nonhalogenated VOCs
Halogenated VOCs
Halogenated VOCs
BTEX; Nonhalogenated
VOCs
Status as of
March 2012
predesign
operating
operating
predesign
operating
completed
predesign
predesign
completed
operating
design
operating
Start Date
Sep 2004
2000
2005
2001
Nov2002
Sep 2002
Jun 2006
End Date
present
present
present
2002
2004
present
present
Introduction to In Situ Bioremediation of Groundwater
A-l
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
HOOKER CHEMICAL &
PLASTICS
CORP./RUCO
POLYMER CORP. - OU
3
ICELAND COIN
LAUNDRY AREA GW
PLUME - Former
Facility Area
ICELAND COIN
LAUNDRY AREA GW
PLUME -Plume Area
MONITOR DEVICES,
INC./INTERCIRCUITS,
INC.
NEPERA CHEMICAL
CO., INC.
ABERDEEN PROVING
GROUND
(MICHAELSVILLE
LANDFILL) -Site 16
DRMO Metal Scrap
Yard
ABERDEEN PROVING
GROUND
(MICHAELSVILLE
LANDFILL) -Site 23
Building 525 Site
ABERDEEN PROVING
GROUND
(MICHAELSVILLE
LANDFILL)- Site 28f-
Building3327UST
Site
OU
03
01
01
01
01
06
06
06
Document
Type
ROD
ROD
ROD
ROD
ROD
ROD
ROD
EPA ID
NYD002920312
NJ0001360882
NJ0001360882
NJD980529408
NYD000511451
MD3210021355
MD3210021355
MD3210021355
EPA
Region
02
02
02
02
02
03
03
03
Bioremediation
Type
Biosparging
In situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
Biosparging
Biobarrier
Biobarrier
Direct injection
N/A
Direct injection
Direct injection,
Recirculation
Contaminants
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
BTEX; Halogenated
VOCs; Nonhalogenated
SVOCs;
Nonhalogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Status as of
March 2012
operating
operating
operating
operating
being
installed
operating
operating
operating
Start Date
Oct 2006
May 2007
Apr 2007
2010
Fall 2011
2007
2006
2006
End Date
present
present
present
present
present
present
present
Introduction to In Situ Bioremediation of Groundwater
A-2
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
ABERDEEN PROVING
GROUND
(MICHAELSVILLE
LANDFILL) -Site 29
Tower Road Site
ABERDEEN PROVING
GROUND
(MICHAELSVILLE
LANDFILL) -Site 32
Building 507 Site
ABERDEEN PROVING
GROUND
(MICHAELSVILLE
LANDFILL) -Site 33
Building MSOOSite
ANDREWS AIR FORCE
BASE - FT-04
ANDREWS AIR FORCE
BASE - ST-10 (PD-680
Spill)
ANDREWS AIR FORCE
BASE - ST-14 (East
Side Gas Station)
Benzene Plume
ANDREWS AIR FORCE
BASE - ST-14 (East
Side Gas Station) TCE
and TCE/CT Plumes
AVCOLYCOMING
(WILLIAMSPORT
DIVISION) -Shallow
Aquifer
BRANDYWINEDRMO
OU
06
06
06
03
07
11
11
02
01
Document
Type
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
EPA ID
MD3210021355
MD3210021355
MD3210021355
MD0570024000
MD0570024000
MD0570024000
MD0570024000
PAD003053709
MD9570024803
EPA
Region
03
03
03
03
03
03
03
03
03
Bioremediation
Type
Bioaugmentation, In
situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
Biobarrier
Recirculation
Biobarrier,
Direct injection,
Recirculation
Direct injection
Direct injection
Biobarrier
Biobarrier
Direct injection
Direct injection
Contaminants
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
BTEX; Halogenated
VOCs; Nonhalogenated
VOCs
BTEX; Nonhalogenated
SVOCs;
Nonhalogenated VOCs
BTEX; Nonhalogenated
VOCs
Halogenated VOCs
Metals and metalloids
Halogenated VOCs
Status as of
March 2012
operating
operating
operating
operating
operating
operating
operating
completed
operating
Start Date
2007
2006
2006
Aug 2004
Sep 2004
May 2006
May 2006
1997
Feb 2008
End Date
present
present
present
present
present
present
present
2000
present
Introduction to In Situ Bioremediation of Groundwater
A-3
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
BRESLUBE-PENN, INC.
DEFENSE GENERAL
SUPPLY CENTER (DLA)
DOVER AIR FORCE
BASE- Area 2 Plume
DOVER AIR FORCE
BASE -Area 5 Plume
DOVER AIR FORCE
BASE- Area 6 Plume
DOVER AIR FORCE
BASE-LF25 Plume
DOVER AIR FORCE
BASE-OT41/Building
719 Source Zone -
Ongoing Interim
Remedy
DOVER AIR FORCE
BASE -SS08 Plume
FIKE CHEMICAL, INC.
INDIAN HEAD NAVAL
SURFACE WARFARE
CENTER - Site 57
Building 292 TCE
Contamination
Downgradient Plume
OU
01
08
15
17
16
19
16
19
04
01
Document
Type
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD-A
ROD
EPA ID
PAD089667695
VA3971520751
DE8570024010
DE8570024010
DE8570024010
DE8570024010
DE8570024010
DE8570024010
WVD047989207
MD7170024684
EPA
Region
03
03
03
03
03
03
03
03
03
03
Bioremediation
Type
Bioremediation
Unknown Type
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Biosparging
In situ Aerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
N/A
N/A
Direct injection
Direct injection,
Recirculation
Direct injection,
Recirculation
Direct injection,
Recirculation
Direct injection,
Recirculation
Direct injection,
Recirculation
Biosparging
N/A
Contaminants
BTEX; Halogenated
VOCs; Nonhalogenated
VOCs
Halogenated VOCs
Halogenated VOCs
BTEX; Halogenated
VOCs; Nonhalogenated
VOCs
BTEX; Halogenated
SVOCs; Halogenated
VOCs; Nonhalogenated
VOCs; Organic
pesticides
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
BTEX; Halogenated
SVOCs; Halogenated
VOCs; Metals and
metalloids;
Nonhalogenated
SVOCs;
Nonhalogenated VOCs;
Organic pesticides
Halogenated VOCs
Status as of
March 2012
design
predesign
operating
operating
operating
operating
operating
operating
operating
designed/not
installed
Start Date
2006
2006
2006
2006
2002
2006
Jun 2007
End Date
present
Introduction to In Situ Bioremediation of Groundwater
A-4
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
INDIAN HEAD NAVAL
SURFACE WARFARE
CENTER - Site 57
Building 292 TCE
Contamination
Source Zone
LETTERKENNYARMY
DEPOT (SE AREA)
MARINE CORPS
COMBAT
DEVELOPMENT
COMMAND
NAVAL AMPHIBIOUS
BASE LITTLE CREEK -
Site 11 Plating Shop
NAVAL AMPHIBIOUS
BASE LITTLE CREEK -
Site 12 Exchange
Laundry
NAVAL AMPHIBIOUS
BASE LITTLE CREEK -
Site 13 PCP Tank
NAVAL SURFACE
WARFARE CENTER -
DAHLGREN-Site20A
Plume
NAVAL SURFACE
WARFARE CENTER -
DAHLGREN-Site20B
Plume
NAVAL SURFACE
WARFARE CENTER -
DAHLGREN-Site23
Plume
OU
01
10
19
05
06
07
19
19
19
Document
Type
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
EPA ID
MD7170024684
PA6213820503
VA1170024722
VA5170022482
VA5170022482
VA5170022482
VA7170024684
VA7170024684
VA7170024684
EPA
Region
03
03
03
03
03
03
03
03
03
Bioremediation
Type
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
Direct injection
Direct injection
N/A
Biobarrier, Direct
injection
Biobarrier
Biobarrier
Direct injection
Direct injection
Direct injection
Contaminants
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated SVOCs;
Halogenated VOCs;
Organic pesticides
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Status as of
March 2012
operating
operating
predesign
operating
operating
operating
operating
operating
operating
Start Date
Dec 2011
1999
Apr 2009
Mar 2007
May 2010
2009
2009
2009
End Date
present
Jun 2007
present
present
present
Introduction to In Situ Bioremediation of Groundwater
A-5
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
^H
PATUXENT RIVER
NAVAL AIR STATION -
Site 39 Waste PCE
Storage Area
(Building 503)
SAEGERTOWN
INDUSTRIAL AREA
SAND, GRAVEL AND
STONE -OU3-
Shallow GW in
Eastern Excavation
Area
CAPE FEAR WOOD
PRESERVING
DISTLER BRICKYARD -
Bioremediation
ESCAMBIAWOOD-
PENSACOLA-High
Concentration Plume
Areas
ESCAMBIAWOOD-
PENSACOLA - Source
Plume Area
PCX, INC.
(STATESVILLE PLANT)
JACKSONVILLE NAVAL
AIR STATION -OU3-
Area C Hot Spot
JACKSONVILLE NAVAL
AIR STATION - OU3 -
Area D Hot Spot
OU
24
01
03
01
01
02
02
03
03
03
Document
Type
ROD
ROD-A
ROD
ROD-A
FYR
ROD
ROD
ESD
ROD
ROD
EPA ID
MD7170024536
PAD980692487
MDD980705164
NCD003188828
KYD980602155
FLD008168346
FLD008168346
NCD095458527
FL6170024412
FL6170024412
EPA
Region
03
03
03
04
04
04
04
04
04
04
Bioremediation
Type
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
Direct injection
Biobarrier, Direct
injection
N/A
Direct injection,
Recirculation
Direct injection,
Hydraulic fracturing
N/A
N/A
Biobarrier
Direct injection
Direct injection
Contaminants
Halogenated VOCs
Halogenated VOCs
BTEX; Halogenated
VOCs; Nonhalogenated
VOCs
Nonhalogenated
SVOCs
Halogenated VOCs
Halogenated SVOCs;
Nonhalogenated
SVOCs;
Nonhalogenated VOCs;
Organic pesticides
Halogenated SVOCs;
Nonhalogenated
SVOCs;
Nonhalogenated VOCs;
Organic pesticides
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Status as of
March 2012
operating
operating
design
completed
operating
design
design
operating
operating
operating
Start Date
Oct 2009
2003/2004
Aug 2001
Apr 2003
May 2007
Feb 2003
Dec 2002
End Date
present
present
Sep 2004
present
present
present
present
Introduction to In Situ Bioremediation of Groundwater
A-6
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
JACKSONVILLE NAVAL
AIR STATION -OU5
LANDIA CHEMICAL
COMPANY -Interim
Remedy- Operable
Unit 2 Groundwater
MEMPHIS DEFENSE
DEPOT (DLA) Main
Installation Functional
Unit7-TTA-land2
OAK RIDGE
RESERVATION
(USDOE) - East Bethel
Valley VOC Plume
(7000-Area)
PALMETTO WOOD
PRESERVING
PEAK OIL CO. /BAY
DRUM CO. -Surficial
Aquifer
PICAYUNE WOOD
TREATING SITE
TOWER CHEMICAL
CO.
ou
05
02
2/3/4
30
01
02
00
03
Document
Type
ROD
ROD
ROD-A
ROD-A
ROD
ROD
EPA ID
FL6170024412
FLD042110841
TN4210020570
TN 1890090003
SCD003362217
FLD004091807
MSD065490930
FLD004065546
EPA
Region
04
04
04
04
04
04
04
04
Bioremediation
Type
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Bioremediation
Bioaugmentation, In
situ Aerobic
Bioremediation
Bioaugmentation, In
situ Aerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
N/A
N/A
Biobarrier
N/A
Biobarrier
Pneumatic
fracturing
Direct injection
N/A
N/A
Contaminants
BTEX; Halogenated
VOCs; Nonhalogenated
SVOCs;
Nonhalogenated VOCs
Nitrate
Halogenated VOCs
Halogenated VOCs
Metals and metalloids
Halogenated VOCs
Halogenated SVOCs;
Nonhalogenated
SVOCs;
Nonhalogenated VOCs;
Organic pesticides
BTEX; Halogenated
SVOCs; Halogenated
VOCs; Metals and
metalloids;
Nonhalogenated
SVOCs;
Nonhalogenated VOCs;
Organic pesticides
Status as of
March 2012
predesign
predesign
operating
design
operating
operating
designed/not
installed
predesign
Start Date
Sep 2006
Jan 2009
Jun 2005
End Date
present
present
present
Introduction to In Situ Bioremediation of Groundwater
A-7
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
^H
USN AIR STATION
CECIL FIELD -Site 59
(Hot Spot Nos. 2 and
3)
AIRCRAFT
COMPONENTS (D & L
SALES) Chemical
Operable Unit OU-2
GALESBURG/KOPPERS
CO.- Deep sand
aquifer
GALESBURG/KOPPERS
CO. -Shallow till
aquifer
KOPPERSCOKE-
Groundwater OU
PARSONS CASKET
HARDWARE CO. -
Alluvial Aquifer
PARSONS CASKET
HARDWARE CO. -
Bedrock Groundwater
Aquifer
TAR LAKE - OU2
AMERICAN CREOSOTE
WORKS, INC.
(WINNFIELD PLANT)
GRANTS
CHLORINATED
SOLVENTS
OU
09
02
01
01
01
02
02
02
01
00
Document
Type
ROD
ESD
ESD
ROD
ROD
ROD
ROD
ROD
EPA ID
FL5170022474
MI0001119106
ILD990817991
ILD990817991
MND000819359
ILD005252432
ILD005252432
MID980794655
LAD000239814
NM0007271768
EPA
Region
04
05
05
05
05
05
05
05
06
06
Bioremediation
Type
Bioaugmentation, In
situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Aerobic
Bioremediation
Biosparging
Bioaugmentation, In
situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
Biosparging
Bioaugmentation, In
situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
Recirculation
Direct injection
Recirculation
Recirculation
Biosparging
N/A
N/A
Biosparging
Direct injection,
Recirculation
Biobarrier
Contaminants
Halogenated VOCs
Halogenated VOCs
Halogenated SVOCs;
Nonhalogenated
SVOCs;
Nonhalogenated VOCs;
Organic pesticides
Halogenated VOCs;
Nonhalogenated
SVOCs;
Nonhalogenated VOCs
BTEX; Nonhalogenated
SVOCs;
Nonhalogenated VOCs
Halogenated VOCs
Halogenated VOCs
BTEX; Nonhalogenated
VOCs
BTEX; Halogenated
SVOCs;
Nonhalogenated
SVOCs;
Nonhalogenated VOCs;
Organic pesticides
Halogenated VOCs
Status as of
March 2012
operating
operating
operating
operating
completed
predesign
predesign
operating
operating
operating
Start Date
2008
Aug 2004
1998
1998
Oct 1996
Dec 2010
End Date
present
present
present
present
1999
present
present
present
Introduction to In Situ Bioremediation of Groundwater
A-8
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
^•_
MCGAFFEYAND
MAIN
GROUNDWATER
PLUME -Hotspot in
Groundwater Plume
Area
NORTH RAILROAD
AVENUE PLUME -
Deep Zone
NORTH RAILROAD
AVENUE PLUME -
Downgradient
Biocurtain
NORTH RAILROAD
AVENUE PLUME -
Source Area and
Hotspot
OUACHITA NEVADA
WOOD TREATER
PANTEX PLANT
(USDOE)- Southeast
Area ISB System
PANTEX PLANT
(USDOE) -Zone 11 ISB
System
PETRO-CHEMICAL
SYSTEMS, INC.
(TURTLE BAYOU) -
Shallow Groundwater
SOL
LYNN/INDUSTRIAL
TRANSFORMERS -
Bioremediation
HASTINGS GROUND
WATER
CONTAMINATION -
Far-Mar-Co Subsite
OU
00
01
01
01
01
00
00
02
02
06
Document
Type
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD-A
ROD-A
ROD
EPA ID
NM0000605386
NMD986670156
NMD986670156
NMD986670156
ARD042755231
TX4890110527
TX4890110527
TXD980873350
TXD980873327
NED980862668
EPA
Region
06
06
06
06
06
06
06
06
06
07
Bioremediation
Type
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Unknown Type
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Aerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
N/A
Direct injection
Biobarrier,
Recirculation
Recirculation
N/A
Biobarrier
Biobarrier
Recirculation
Biobarrier, Direct
injection
Direct injection
Contaminants
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated SVOCs;
Organic pesticides
Explosives/propellants;
Metals and metalloids;
Nonhalogenated
SVOCs
Explosives/propellants;
Halogenated VOCs
BTEX; Halogenated
VOCs; Nonhalogenated
SVOCs;
Nonhalogenated VOCs
Halogenated VOCs
Halogenated VOCs
Status as of
March 2012
design
operating
operating
operating
predesign
operating
operating
completed
operating
operating
Start Date
Apr 2008
May 2008
May 2008
Feb/Mar
2008
Jun 2009
1997
2010
Jul2010
End Date
present
present
present
present
present
2005
present
present
Introduction to In Situ Bioremediation of Groundwater
A-9
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
HASTINGS GROUND
WATER
CONTAMINATION -
Second Street OU
IOWA ARMY
AMMUNITION PLANT
- Off-site
Groundwater
LAKE CITY ARMY
AMMUNITION PLANT
(NORTHWEST
LAGOON) - Area 12
Groundwater
LAKE CITY ARMY
AMMUNITION PLANT
(NORTHWEST
LAGOON) - Area 18
Paleochannels
LAKE CITY ARMY
AMMUNITION PLANT
(NORTHWEST
LAGOON) - Area 18
Shallow VOC Source
Areas
LAKE CITY ARMY
AMMUNITION PLANT
(NORTHWEST
LAGOON) -NE Corner
Operable Unit Area
16B Plume
LAKE CITY ARMY
AMMUNITION PLANT
(NORTHWEST
LAGOON) -NE Corner
Operable Unit Area
17B Downgradient
Plume (IRZ Line 5)
OU
20
03
01
02
02
03
03
Document
Type
ROD
ROD
ROD
ROD
ROD
ROD
ROD
EPA ID
NED980862668
IA7213820445
M03213890012
M03213890012
M03213890012
M03213890012
M03213890012
EPA
Region
07
07
07
07
07
07
07
Bioremediation
Type
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
Biobarrier
Biobarrier
Biobarrier
Biobarrier
N/A
Biobarrier
Biobarrier
Contaminants
BTEX; Nonhalogenated
SVOCs;
Nonhalogenated VOCs
Explosives/propellants;
Nonhalogenated
SVOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Status as of
March 2012
operating
operating
operating
operating
installed
operating
operating
Start Date
Nov2005
Oct 2007
Feb 2008
Oct 2007
Jan 2008
Oct 2007
End Date
present
present
present
present
present
present
Introduction to In Situ Bioremediation of Groundwater
A-10
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
LAKE CITY ARMY
AMMUNITION PLANT
(NORTHWEST
LAGOON) -NE Corner
Operable Unit Area
17B Source Area
Residual NAPLZone
(IRZ Lines 1-4)
LAKE CITY ARMY
AMMUNITION PLANT
(NORTHWEST
LAGOON) -NE Corner
Operable Unit Area
17D Plume
MISSOURI ELECTRIC
WORKS -Alluvial
Groundwater
BOUNTIFUL/WOODS
CROSS 5TH S. PCE
PLUME
BOUNTIFUL/WOODS
CROSS 5TH S. PCE
PLUME
F.E. WARREN AIR
FORCE BASE -Spill
Site 7 (SS-7) Plume
IDAHO POLE CO.
LIBBY GROUND
WATER
CONTAMINATION -
Boundary Injection
System
OU
03
03
02
01
02
02
01
02
Document
Type
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
EPA ID
M03213890012
M03213890012
MOD980965982
UT0001119296
UT0001119296
WY5571924179
MTD006232276
MTD980502736
EPA
Region
07
07
07
08
08
08
08
08
Bioremediation
Type
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
Bioaugmentation, In
situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Aerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
Biobarrier
Biobarrier, Direct
injection
N/A
Biobarrier
Direct injection,
Direct injection,
Hydraulic fracturing
Recirculation
Recirculation
Contaminants
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated SVOCs;
Nonhalogenated
SVOCs; Organic
pesticides
Halogenated SVOCs;
Nonhalogenated
SVOCs; Organic
pesticides
Status as of
March 2012
operating
operating
predesign
operating
operating
operating
operating
completed
Start Date
Oct 2007
Jan 2008
Jul2011
Feb 2011
approx.
2008
1997
1993
End Date
present
present
present
present
2003
Introduction to In Situ Bioremediation of Groundwater
A-ll
-------
Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
LIBBY GROUND
WATER
CONTAMINATION -
Intermediate Aquifer
LOCKWOOD SOLVENT
GROUND WATER
PLUME -Beall
Property
LOCKWOOD SOLVENT
GROUND WATER
PLUME -Plume
Leading Edges
LOCKWOOD SOLVENT
GROUND WATER
PLUME -SOCO
Property
MONTANA POLE AND
TREATING -
Groundwater OU
ALAMEDA NAVAL AIR
STATION - IRSite 16
ALAMEDA NAVAL AIR
STATION -IRSite 6
ALAMEDA NAVAL AIR
STATION -Site 25
Groundwater (Navy
OU5/FISCA IR-02)
ALAMEDA NAVAL AIR
STATION -Site 26
Western Hanger Zone
FORT ORD - FT-044
Operable Unit Carbon
Tetrachloride Plume,
A- Aquifer
FRONTIER FERTILIZER
OU
02
01
01
01
01
01
01
14
06
12
01
Document
Type
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
ROD
EPA ID
MTD980502736
MT0007623052
MT0007623052
MT0007623052
MTD006230635
CA2170023236
CA2170023236
CA2170023236
CA2170023236
CA72 10020676
CAD071530380
EPA
Region
08
08
08
08
08
09
09
09
09
09
09
Bioremediation
Type
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Aerobic
Bioremediation
Unknown Type
Unknown Type
Bioaugmentation,
Biosparging
Bioaugmentation, In
situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
Recirculation
N/A
N/A
N/A
Recirculation
N/A
N/A
Biosparging
Recirculation
Recirculation
N/A
Contaminants
Halogenated SVOCs;
Nonhalogenated
SVOCs; Organic
pesticides
Halogenated VOCs
Halogenated VOCs
Halogenated VOCs
Halogenated SVOCs;
Organic pesticides
Halogenated SVOCs;
Halogenated VOCs;
Organic pesticides
Halogenated VOCs
BTEX; Nonhalogenated
SVOCs;
Nonhalogenated VOCs
Halogenated VOCs
Halogenated VOCs
Nitrate
Status as of
March 2012
completed
predesign
predesign
predesign
completed
predesign
predesign
operating
operating
operating
predesign
Start Date
1993 (?)
1999/2000
Mar 2009
Sep 2010
Sep 2009
End Date
1998
2002
present
present
present
Introduction to In Situ Bioremediation of Groundwater
A-12
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Appendix A: A Selection of Superfund Program In situ Groundwater Bioremediation Sites
(Remedies Selected FY1989 to 2008)
Site Name
KOPPERS CO INC
(OROVILLE PLANT) -
Off-Property Plume
KOPPERS CO., INC.
(OROVILLE PLANT) -
On-Property East
Plume
SELMA TREATING CO.
IDAHO NATIONAL
ENGINEERING
LABORATORY
(USDOE) Test Area
North OU 1-07B
(OU1) hot spot
OU
01
01
01
01
Document
Type
ROD-A
ROD-A
ESD
ROD-A
EPA ID
CAD009112087
CAD009112087
CAD029452141
ID4890008952
EPA
Region
09
09
09
10
Bioremediation
Type
In situ Aerobic
Bioremediation
In situ Aerobic
Bioremediation
In situ Anaerobic
Bioremediation
In situ Anaerobic
Bioremediation
Treatment
Approach
(Operating &
Completed Projects
Only)
Direct injection
Direct injection
Direct injection,
Recirculation
Direct injection
Contaminants
Halogenated SVOCs;
Organic pesticides
Halogenated SVOCs;
Organic pesticides
Metals and metalloids
Halogenated VOCs
Status as of
March 2012
operating
operating
operating
operating
Start Date
Aug 1998
Mar 1998
Mar 2005
1999
End Date
present
present
present
present
Abbreviations/Acronyms:
BTEX = Benzene, Toluene, Ethyl Benzene, and Xylene
ESD = Explanation of Significant Differences
FY = Fiscal Year
FYR = Five-Year Review
ISB = In situ Bioremediation
N/A = Not Applicable
OU = Operable Unit
ROD = Record of Decision
ROD-A = Record of Decision Amendment
TBD = To Be Determined
SVOC = Semivolatile Organic Compound
VOC = Volatile Organic Compound
Introduction to In Situ Bioremediation of Groundwater
A-13
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