EPA/600/R-13/237
September 2013
xvEPA
United States
Environmental Protection
Agency
Ground Water Issue
Ground Water Issue Paper: Synthesis Report on State of
Understanding of Chlorinated Solvent Transformation
Bruce Pivetz, Ann Keeley*, Eric Weber, Jim Weaver, John Wilson, and Cissy Ma
Contents
1. Objectives and Scope 1
2. Introduction 2
2.1 MNA 2
2.2 Site Characterization and Conceptual Site Model. 2
2.3 Physical and Chemical Properties of the Contaminants. 2
2.4 Contaminant Transport and Physical Attenuation Processes 4
2.5 Geochemical Conditions. 5
2.6 Contaminant Natural Attenuation Rates 6
3. Biotic Chlorinated Solvent Transformation Pathways
and Processes 7
3.1 Introduction to BioticTransformations 7
3.2 PCEandTCE 9
3.3 TCA 24
3.4 Dioxane 25
4. Abiotic Transformations 27
4.1 PCEandTCE 29
4.2 TCA 33
4.3 Dioxane 35
5. Summary of Biotic and Abiotic Transformations 35
6. Modeling applications and conceptualizations for
chlorinated solvent transformations 37
6.1 Historical Background 37
6.2 Types of Models 37
6.3 Parameter Measurement in the Field 38
6.4 Model Application 38
7. References 41
Figures
Figure 1.1. Elements of a conceptual site model for monitored
natural attenuation. 3
Figure 3.1. Bacterial species involved in dechlorination
processes. 23
Figure 3.2. Enzymes involved in dechlorination processes. 23
Figure 4.1. Formation of abiotic reductants as a function of iron
and sulphate reducing zones. 29
Figure 4.2. Reaction Scheme illustrating the degradation
pathways for PCE. 30
Figure 4.3. Reaction scheme illustrating the degradation
pathways for TCA 34
Tables
Table 1. Contaminant physical and chemical properties. 4
Table 2a. Microbial Metabolic Processes 8
Table 2b. Reactions and Subsurface Conditions 8
Table 3a. Compilation of compilations of chlorinated solvent 11
Table 3b. Chlorinated solvent biotic transformation 21
Table 4.1. Surface area-normalized rate constants 31
* Corresponding author: National Risk Management Research
Laboratory, U.S. Environmental Protection Agency, 919
Kerr Research Drive, Ada, OK 74820, USA
Tel.: 1.580.436.8890 fax: 1.580.436.8614
Email: keeley.ann@epa.gov (A. Keeley)
1. OBJECTIVES AND SCOPE
Chlorinated solvents are altered by biotic and abiotic processes.
Biotic transformation can include reductive dechlorination,
cometabolism, and limited oxidation. Abiotic transformation
is less well understood but may play a role at some sites.
Transformations may be limited such that endpoints fall short of
complete degradation of the solvent to innocuous compounds.
Determination of which endpoints are reached, the processes of
transformation, and the needed site data are critical for assessing
and modeling transport, and deciding on monitored natural
attenuation (MNA) as a remedy.
This Issue Paper summarizes the biotic and abiotic transformations
of several important chlorinated solvents. It briefly describes the
factors that affect the transformation mechanisms, as well as the
measurements necessary to distinguish among the mechanisms. It
serves as a guide for developing an advanced ground-water transport
model, with governing equations for simulating these processes
in models. The primary audience is the EPA remedial project
managers (RPMs). The Issue Paper is intended to provide RPMs
with a basic understanding of the fundamentals and terminology of
chlorinated solvent transformation in the context of MNA.
The focus of this document is on three chlorinated solvents
used at industrial and dry-cleaning facilities: tetrachloroethene
(PCE), trichloroethene (TCE), and 1,1,1-trichloroethane
(TCA). It also discusses their degradation ("daughter") products:
1,2-dichloroethene (DCE) [primarily r»-l,2-dichloroethene (cis-
DCE)], vinyl chloride (VC), 1,1-dichloroethene (1,1-DCE),
1,1-dichloroethane (1,1-DCA), and chloroethane (CA). It also
covers 1,4-dioxane (dioxane), which is present as a stabilizer in
some chlorinated solvent preparations [it was primarily used to
stabilize TCA (Mohr 2001)]. These chlorinated solvents are among
the most commonly encountered contaminants at many of the
worst contaminated sites, and PCE is the primary contaminant
found at dry-cleaner sites. TCE is also found at dry-cleaner sites
as a degradation product of PCE, and as the initial contaminant at
older dry cleaning sites as it was the dry-cleaning agent used for a
few decades starting about 1930.
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2. INTRODUCTION
Understanding and modeling the fate (transformation)
and transport of chlorinated solvents at contaminated
sites, as well as their remediation through the
use of MNA, requires a thorough recognition of
transformation processes to form a strong foundation
for conceptual modeling. This introductory section
presents brief discussions of these topics.
2.1 MNA
The U.S. EPA (1999) provided clarification of its
policy on the application of MNA as a remedy for
contaminated sites, and defines this alternative as
"the reliance on natural attenuation processes (within
the context of a carefully controlled and monitored site
cleanup approach) to achieve site-specific remediation
objectives within a time frame that is reasonable
compared to that offered by other more active methods".
Natural attenuation (NA) processes that degrade
or destroy contaminants are preferred over other
processes (e.g., dilution and volatilization) that merely
attenuate (i.e., diminish contaminant concentrations)
contaminant mass (U.S. EPA, 1999). By definition,
MNA does not include the use of any active remedial
technologies; however, at most sites MNA is very
likely to be just one component of the overall
remedial strategy as it may be applied to only certain
portions of the site, and/or after active technologies
have been implemented. Thus, when investigating,
modeling, or evaluating MNA it is imperative to
take into consideration other remedial activities that
have previously occurred or are currently taking
place. Guidance documents have been used for
implementing MNA in ground water (Wiedemeier et
al., 1998, 1999; U.S. EPA, 1999; National Research
Council, 2000).
2.2 Site Characterization and Conceptual
Site Model.
Information about the subsurface contamination,
geology, hydrogeology, geochemistry, and microbiology
collected during site characterization is assembled
into a conceptual site model (Figure 1.1). The
conceptual site model (CSM) is "a three-dimensional
representation that conveys what is known or suspected
about contamination sources, release mechanisms, and
the transport and fate of those contaminants" (U.S. EPA,
1999). The elements of site characterization and
the process of preparing a CSM for MNA of volatile
organic compounds (VOCs; including chlorinated
solvents) are described in Pivetz et al. (2012).
Defining the plume in three dimensions and
understanding the geochemical and microbiological
environment are necessary parts of establishing the
CSM. Identifying and defining the most significant
ground-water and contaminant flow path(s), and
quantifying flow velocities, are critical for estimating
chlorinated solvent attenuation rates. Characterization
of the subsurface geochemistry is also important.
Microbiological characterization and confirmation
of the presence of specific bacterial strains is likely to
be important to fully evaluate MNA for PCE, TCE,
TCA, and dioxane, since effective bioattenuation
of each of these depends on the presence of specific
microbes. Monitoring should be extensive enough
in three dimensions to be able to understand the
differing conditions that are likely to occur in different
portions of the site and plume. Monitoring should
be conducted for a long enough period (likely several
years) in order to estimate rates of attenuation at a
given location.
Development of the CSM and modeling of the
plume migration, attenuation, and duration requires
knowledge of physical characteristics of the subsurface,
and activities and changes at the site. According to
Pivetz et al. (2012), information related to the ground-
water and contaminant velocities, the lithology (which
impacts contaminant transport and sorption such as
back-diffusion), seasonal changes impacting ground-
water levels, longer-term changes (e.g., droughts), and
the role and impacts of active remedial technologies
(especially source removal activities) should be
collected during site characterization.
2.3 Physical and Chemical Properties of the
Contaminants.
Table 1 presents the most significant physical
and chemical properties, and their values, of the
chlorinated solvents that impact the fate and transport
of these compounds in the subsurface. One set of
values is provided in the table; however, it should
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Contaminant Source and Source
Control Information
• Location, nature, and history of contaminant releases
or sources
• Locations and characlerizalions of sources for
ground-water contamination [e.g., nonaqueous
phase liquid (NAPL)]
• Locations and descriptions of source control and
other ongoing and proposed remedial actions
Geologic and Hydrologic Information
1 Regional and site geologic and hydrologic settings.
including controls on ground-water flow
• Analyses of depositional environments and geologic
features that may serve as zones of preferential flow or
barriers to flow, including geometry and physical
properties of geologic facies (e.g., texture, porosity, bulk
density) and their variability
•Stratigraphy, including thickness and lateral continuity
of geologic units, and bedding features
'Anthropogenic features (e.g., buried corridors and
heterogeneous fill materials) that control ground-water
flow, and may serve as migration pathways or barriers
1 Depth lo ground water and temporal variation
• Characteristics of surface water bodies (e.g.. locations,
depths, and flow rates), their interactions with ground
water, and temporal variations
1 Ground-water recharge and discharge locations, rates
and temporal variability
• Hydraulic gradients, including horizontal and vertical
components, and their variations in response to
fluctuations in site hydrology (e.g., seasonal or longer
term precipitation patterns and changes in patterns of
ground-water withdrawal or irrigation)
p Hydraulic properties (e.g., hydraulic conductivities,
storage properties, and effective porosities) and their
variability and anisotropy within geologic units
• Quantitative description of the ground-water flow field
p Chemical properties of the subsurface matrix including
mineralogy and organic matter
Receptor Information
• Aquifer classification, current usage information, and
reasonably anticipated future usage
• Locations and production data for water-supply wells
• Locations and information on human and ecological
receptors under current and reasonably anticipated
future conditions
•Areas susceptible to impact by vapor-phase
contaminants (e.g.. indoor air)
1 Information on local historical and cultural uses of land,
water, and older resources used lo identify receptor
populations
• Descriptions of institutional controls currently in place
Contaminant Distribution, Transport and Fate
• Distribution of each contaminant phase (i.e., gaseous,
aqueous, sorbed, NAPL) and estimates of mass
• Mobility of contaminants in each phase
• Temporal trends in contaminant mass and
concentrations
• Sorption information, Including retardation factors,
sorption mechanics, and controls
• Contaminant attenuation processes and rate estimates
• Assessment of facilitated transport mechanisms (e.g.,
complexation or colloidal transport)
• Geochemieal characteristics that affect or are indicative
of contaminant transport and fate, and mineralogy, if
needed
1 Potential for mobilization of secondary contaminants
(e.g., arsenic)
• Effects of other proposed or ongoing remedial
activities on contaminant transport, fate, and natural
attenuation processes
Figure 1.1. Elements of a conceptual site model for monitored natural attenuation.
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatic
jnd Water Issue 3
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Table 1. Contaminant physical and chemical properties.
Properties:
Molecular
weight
(g mole"1)
Water
solubility
at 25 °C
(mgL-1)
Contaminant
density
as a NAPL
(gmL-1)
Log soil/water
partition
coefficient
(Log Kow)
Vapor
pressure
at 20 or 25 °C
[mm Hg)
Henry's Law
Constant
at 25 °C
(atm m3 mol"1)
PCE1
C2CI4
(CI2C=CCI2)
165.83
150
1.6227
3.40
18.47
1.8x 10'2
TCE1
C2HCI3
(HCIC=CCI2)
131.4
1070
1.465
2.42
74
1.1 x10'2
Cis-DCE1
C2H2CI2
(HCIC=CCIH)
96.95
3500
1.2837
1.86
180
4.08 x10'3
vc1
C2H3CI
(H2C=CCIH)
62.5
2763
0.9106
1.36
2530
2.78 x10'2
TCA1
CI3CH3
(CCI3CH3)
133.4
1500
1.3390
2.49
124
6.3 x10'3
1,1 -DCA1
C2H4CI2
(HCI2C-CH3)
98.97
5500
1.1747
1.79
1.82
4.4 x10'2
CA1
C2H5CI
(CH3-CH2-CI)
64.52
5740
0.9214
1.43
1008
1.11 X10-2
Dioxane1'2
(C4H802)
88.11
Miscible
1.0329
-0.27
38.1
5x 10'6
References
1ATSDR: Toxicological Profiles for each compound
2Mahendra and Alvarez-Cohen (2006)
be noted that the values of these properties can vary,
depending on the conditions and how the values were
measured.
2.4 Contaminant Transport and Physical
Attenuation Processes
The attenuation of contaminant concentrations
with time and distance from a source area (i.e.,
natural attenuation) can be due to "a variety of
physical, chemical, or biological processes that...
include biodegradation; dispersion; dilution; sorption;
volatilization; radioactive decay; and chemical or
biological stabilization, transformation, or destruction
of contaminants" (U.S. EPA, 1999). Detailed
discussion of these processes can be found in the
following contaminant hydrogeology reference
books: Freeze and Cherry (1979), Fetter (1993),
and Domenico and Schwartz (1998). This
document focuses on the destructive processes:
biotic transformations (biodegradation) and abiotic
transformations (degradation through chemical
reactions). However, confirming and quantifying the
impacts of these destructive processes (and calculation
and understanding of attenuation rates) requires an
understanding of how the other, non-destructive,
processes impact the site and the data collected for
the MNA evaluation. All of the contaminants in this
document are subject to advection, dispersion, and
dilution. Chlorinated solvent concentrations will be
relatively low in ground water due to a low solubility.
Dioxane, however, is miscible with water, meaning
ground-water concentrations can be quite high. This
means that sorption will be negligible for dioxane,
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whereas, the chlorinated solvent contaminants will
be slightly to moderately sorbed. Volatilization from
shallow ground water can occur with the chlorinated
solvents; however, volatilization is unimportant for
dioxane due to its very low Henry's Law constant.
None of these contaminants are subject to radioactive
decay. All of these contaminant fate and transport
processes and properties need to be recognized and
quantified in order to model contaminant ground-
water migration and to quantify the effectiveness
of NA. This includes understanding the relative
significance of each of the processes.
2.5 Geochemical Conditions.
Biotic and abiotic transformations of the chlorinated
solvents will be influenced by the subsurface soil and
ground-water geochemical conditions, which may vary
with time and location. Most significantly, which
oxidation-reduction (redox) reactions occur in the
subsurface will determine whether or not a particular
contaminant is transformed, and the extent and rate
of its transformation. Identification of different zones
of different redox conditions and processes will help
indicate where particular transformations are or are
not occurring. The main redox reactions (terminal e"
accepting processes, or TEAPs), their final or terminal
e" acceptors (TEAs), and their reaction products that
occur or are found in the subsurface are:
• Aerobic respiration: the TEA is oxygen (O2), and
CO2 is produced.
• Nitrate reduction (denitrification): the TEA is
nitrate (NO3"), and N2 is produced.
• Manganese reduction: the TEA is manganese(IV)
(Mn+4), and manganese(II) (Mn+2) is produced.
• Iron(III) reduction: the TEA is iron(III) (Fe+3), and
iron(II) (Fe+2) is produced.
• Sulfate reduction: the TEA is sulfate (SO42"), and
hydrogen sulfide (H2S) is produced.
• Methanogenesis: the TEA is carbon dioxide (CO2),
and methane (CH4) is produced.
The TEAPs generally occur in the order given above.
After dissolved oxygen is depleted and the subsurface
becomes anaerobic, the TEAPs shift to denitrification,
then iron(III) reduction and sulfate reduction, and
ultimately to methanogenesis. However, although one
TEAP may be relatively predominant, several of the
TEAPs may occur simultaneously in close proximity
to each other. The occurrence of any one given TEAP
depends on the supply of the terminal e" acceptor and
the appropriate microbial community.
As discussed below in section 3.1.1, the bacteria
that biodegrade chlorinated solvents obtain their
energy during microbiologically mediated oxidation-
reduction reactions in which electrons transfer
between compounds that act as electron (e") donors
(co-contaminants or naturally occurring carbon)
and e" acceptors (the chlorinated solvents). This
reductive dechlorination (which is a major anaerobic
biodegradation pathway for chlorinated solvents) uses
the chlorinated solvents as e" acceptors. Its occurrence
and rate varies depending on the geochemical
conditions brought about during the TEAPs discussed
above, as well as whether the requisite microbes for
dechlorination use the chlorinated solvents or the
TEAs as their e" acceptors.
The predominant redox condition and zone (i.e.,
correlating to a specific microbial TEAP) can also be
identified through subsurface dissolved hydrogen (H2)
measurements (Lovley et al. 1994), as indicated by the
following ranges:
• Denitrification: <0.1 nM H2
• Iron(III) reduction: 0.2 - 0.8 nM H2
• Sulfate reduction: 1-4 nM H2
• Methanogenesis: 5-20 nM H2
Measurement of the TEAs and/or their reduced
products in ground water can indicate what processes
are occurring. Relevant or potentially important
geochemical parameters include soil total organic
carbon (TOC); dissolved organic carbon (DOC);
oxidation-reduction potential (ORP); dissolved
oxygen; (DO); nitrate; manganese (Mn(II)/Mn(IV);
iron (Fe(II)/Fe(III); sulfate; hydrogen sulfide; carbon
dioxide; (CO2); methane, ethane, and ethene;
dissolved hydrogen; pH; alkalinity; temperature;
conductivity; additional major ions such as Ca2+,
Mg2+, K+, Na+, C1-, CO32-, and HCO3-; minerals
present; and concentrations of metals and
metalloids. Not all these parameters will need to be
measured nor will be useful in many cases. Specific
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 5
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geochemical parameters are discussed below for
individual chlorinated solvents when relevant to the
transformation.
In summary, the biotic transformation of the
chlorinated solvents occurs in the subsurface
geochemical environment developed under several
different terminal e" accepting processes that occur in a
general sequence of aerobic oxidation, denitrification,
iron(III) reduction, sulfate reduction, and then
methanogenesis. However, several TEAPs may be
active at one time within the same general subsurface
volume, the predominant TEAP may shift with time,
or a given TEAP may not occur. The occurrence
and significance of a given TEAP depends on the
availability of the relevant electron acceptor.
2.6 Contaminant Natural Attenuation Rates
Calculating the rate of contaminant attenuation in a
ground-water plume is important for evaluating plume
migration and the time frame to reach a remedial
goal. For MNA, attenuation (and contaminant
biodegradation) is often described as a first-order decay
process (i.e., first-order kinetics; exponential decay):
C(t) = C0eM where C(t) = concentration at time t [M L'3]
C0 = initial concentration [M L"3]
k = rate constant [T"1]
t = time [T]
The rate of degradation is given by:
9C/9t =-kC where 9C/9t is the change in concentration
at time t
The rate constant (k) is a critical parameter in
mathematically modeling fate and transport of a
plume. Rate constants for a given process (e.g.,
biodegradation) are often determined under laboratory
conditions, although in NA it is important to
determine the rate constant under site-specific field
conditions. Rate constant values are sometimes
described in terms of half-lives, since they are related
through:
t1/2 = 0.693/k where t1/2 is the half-life [T]
k = first-order rate constant [T"1]
The overall attenuation rate (i.e., rate constant)
representing all transport and attenuation processes at
a single point, or along the entire migration pathway
of the plume, can be calculated using contaminant
concentration data from a sufficient number of
monitoring wells that are properly located in the
migration pathway of the plume. Attenuation
of the source material must also be understood,
as contaminant influx into the plume from the
source area affects the longevity of the plume. The
biodegradation attenuation rate can also be calculated,
which represents the contaminant destructive loss
due only to biological activity. Further discussion of
attenuation rates and methods for their calculation
are provided in Suarez and Rifai (1999) and Newell
et al. (2002). It should be noted that other kinetic
models (e.g., zero order or second order) may be
used to better describe biodegradation or other
transformations of contaminants. Monod kinetics,
as well as the Michaelis-Menten rate law model, is
often used to describe laboratory biodegradation data,
and a variety of kinetic parameters for these kinetics
are determined. Chapelle et al. (2007) discuss the
mathematical treatment of the biotransformation sink
term and kinetics, including substrate and electron
(e") acceptor utilization as described by Monod
kinetics. Alvarez-Cohen and Speitel (2001) provide a
comprehensive discussion of the kinetics involved in
aerobic cometabolism of chlorinated solvents.
Biodegradation and plume attenuation rates (and rate
constants) have been determined from both laboratory
and field studies at contaminated sites. Literature
compilations of rates and rate constants from
numerous sites often do not provide the entire set of
related geochemical, hydrogeological, microbiological,
and anthropogenic conditions, so it may be difficult
to fully understand the conditions that impacted the
rates. Studies where rates and rate constants have been
calculated at chlorinated solvent sites may not have
been published in the peer-reviewed literature, rather,
in gray literature such as site remediation reports.
Laboratory biodegradation rates should be viewed
with caution, as they generally represent much more
optimum conditions than found in the field.
Modeling the potential for NA processes to
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successfully remediate a chlorinated solvent site
depends strongly on knowledge of the rate of biotic
transformation of the solvent(s). Biodegradation rates
and/or rate constants can be calculated from site-
specific measurements, or estimated using previous
knowledge and experience as reflected in the NA
literature.
3. BIOTIC CHLORINATED SOLVENT
TRANSFORMATION PATHWAYS AND
PROCESSES
3.1 Introduction to Biotic Transformations
3.1.1 Biodegradation
In situ biodegradation of chlorinated solvents (i.e.,
biotic transformations) is due primarily to subsurface
bacteria (fungi-mediated biodegradation that may
occur in the unsaturated zone will not occur in the
saturated conditions of ground water). For growth,
bacteria require a carbon source and energy (as well as
water and mineral nutrients) from a substrate(s) (i.e.,
the compound(s) providing the carbon and/or energy).
Heterotrophic bacteria (the majority of bacteria) that
biodegrade chlorinated solvents obtain their carbon
from either naturally occurring compounds or other
contaminants. The energy is obtained from the
energy released during microbiologically mediated
oxidation-reduction reactions in which electrons
transfer between compounds that act as e" donors and
e" acceptors. The e" acceptors can be dissolved oxygen
(O2), some naturally occurring inorganic compounds
(NO3-, Mn+4, Fe+3, SO42-, CO2), or some chlorinated
solvents. In growth-supporting biodegradation, the
contaminant is used as a primary substrate by the
bacteria. Complete biodegradation of the contaminant
to CO2 is termed mineralization. Contaminants
may also be biodegraded through cometabolism, in
which the degradation is non-growth-supporting for
the bacteria bringing about the transformation (the
degradation of the contaminant occurs as a fortuitous
event as the bacteria use some other substrate and the
appropriate enzymes are induced). It is important to
realize that a transformation of a contaminant to an
end product often involves a number of intermediate
compounds and types of reactions, some of which
may not be identified and/or have short persistence.
Reductive dechlorination of PCE and TCE does not
involve any persistent or significant intermediates
before the daughter products DCE and VC are
formed.
Early research on biodegradation of chlorinated
solvents was published by Vogel et al. (1987), Vogel
and McCarty (1987), Sims et al. (1991), Bouwer
(1993), McCarty and Semprini (1993), and Vogel
(1993). MNA microbial processes were discussed
in Azadpour-Keeley et al. (1999), a comprehensive
examination of MNA of petroleum hydrocarbons
and chlorinated solvents is found in Wiedemeier et
al. (1999), and a comprehensive review of chlorinated
solvent MNA is found in Rifai et al. (2001). More
recent reviews of subsurface biodegradation of
VOCs under intrinsic conditions include Field and
Sierra-Alvarez (2004), Lawrence (2006), Aulenta et
al. (2006), Chapelle et al. (2007), and Bradley and
Chapelle (2010).
The literature frequently group chlorinated solvents
biotransformation in a variety of ways: (a) based
on the chemical reaction involved, (b) whether the
contaminant was reduced or oxidized, (c) whether or
not a chlorine was removed, (d) whether the subsurface
conditions were aerobic or anaerobic, (e) whether the
subsurface conditions were oxidizing or reducing, or
(f) by the microbiological metabolic process involved.
Since the same degradative phenomenon may be
referred to in different ways by different practitioners,
it is useful to review and understand the varied
terminology, as well as the basic microbial processes.
Table 2a indicates the biotic transformations of the
contaminants, categorized by the microbial processes
that occur. Table 2b also indicates these biotic
transformations, but categorized by the reactions that
occur. A detailed discussion of relevant terminology
in provided in Bradley and Chapelle (2010). The
broad term "reductive dechlorination" as commonly
used in MNA literature is usually meant to signify
only the specific microbially-mediated process (via
halorespiration, also known as chlororespiration)
resulting in removal of one chloride ion from the
chlorinated compound under anaerobic (reducing)
conditions and its replacement by a hydrogen atom.
However, as indicated by Table 2a and b, other
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Table 2a. Microbial Metabolic Processes Involved in Biotic Transformations of the Chlorinated Solvents.
A. Contaminant as primary substrate:
Growth-supporting.
1. Halorespiration: Anaerobic (anoxic); reductive
dechlorination driven by H2 as an electron donor;
chlorinated solvent used as electron acceptor;
halogen removed (dehalogenation).
2. Respiration/Oxidation: Contaminant used as
electron donor.
a. Oxic respiration: Direct aerobic oxidation.
Oxygen is the terminal electron acceptor.
b. Anoxic respiration: Direct anaerobic oxidation
Inorganic ions other than oxygen are the terminal
electron acceptors.
B. Cometabolism: Non-growth supporting;
contaminant fortuitously degraded with the presence
of another, primary, substrate.
1. Aerobic cometabolism, Cometabolic
oxidation, or Cooxidation: An oxidation reaction;
not a significant naturally occurring process in the
subsurface.
2. Anaerobic cometabolism: A reductive
dechlorination (for the chlorinated solvents); occurs,
but an uncommon and/or slow occurrence; the more
effective anaerobic reductive dechlorination via
halorespiration is not cometabolic.
PCE
Yes
No
No
No
Yes
TCE
Yes
No
No
Yes
Yes
DCE
Yes
Yes
Yes
Yes
Yes
VC
Yes
Yes
Yes
Yes
Yes
TCA
Yes
No
No
Yes
Yes
1,1 -DCA
Yes
Yes
No
Yes
Yes
CA
No
Yes
No
Yes
No
Dioxane
NA
Yes
Maybe
Yes
No
Table 2b. Reactions and Subsurface Conditions Involved in Biotic Transformations of the Chlorinated Solvents.
A. Aerobic oxidation:
a. Direct aerobic oxidation: Oxic respiration.
Oxygen is the terminal electron acceptor.
b. Indirect aerobic oxidation: Aerobic
cometabolism. An oxidation reaction; not a
significant naturally occurring process in the
subsurface.
B. Anaerobic oxidation: Anoxic respiration.
Inorganic ions other than oxygen are the terminal
electron acceptors.
C. Anaerobic reduction: Reductive dechlorination
1. Halorespiration: Anaerobic (anoxic); reductive
dechlorination driven by H2 as an electron donor;
chlorinated solvent used as electron acceptor;
halogen removed; growth-supporting.
2. Anaerobic cometabolism: A reductive
dechlorination, but an uncommon and/or slow
occurrence; the more effective anaerobic reductive
dechlorination via halorespiration is not cometabolic
PCE
No
No
No
Yes
Yes
TCE
No
Yes
No
Yes
Yes
DCE
Yes
Yes
Yes
Yes
Yes
VC
Yes
Yes
Yes
Yes
Yes
TCA
No
Yes
No
Yes
Yes
1,1 -DCA
Yes
Yes
No
Yes
Yes
CA
Yes
Yes
No
No
No
Dioxane
Yes
No
Maybe
NA
NA
8 Ground Water Issu
mthesis Report on State of Understanding of Chlorinated Solvent Transformation
-------
3. Hydrogenolysis: A biotic and abiotic anaerobic
reductive reaction; substitution of a hydrogen
atom for chlorine on the molecule; a reductive
dechlorination; halogens removed (for chlorinated
solvents). When hydrogenolysis is thought of in
terms of being strictly an abiotic reaction, it is likely,
however, to depend on the presence of microbes
to create the conditions conducive to the reaction
(Wiedemeier et al., 1999).
3a. Biotic hydrogenolysis
3b. Abiotic hydrogenolysis
4. Dihaloelimination (dichloroelimination): An
anaerobic reductive reaction; removal of two
adjacent halogen atoms, leaving a double bond
between the respective carbon atoms (forming
an alkene from an alkane); halogens removed
(dehalogenation); a reductive dechlorination. When
dihaloelimination is thought of in terms of being
strictly an abiotic reaction, it is likely, however, to
depend on the presence of microbes to create the
conditions conducive to the reaction (Wiedemeier et
al., 1999).
Yes
Maybe
Maybe
Yes
Maybe
Maybe
Yes
Yes
Yes
Maybe
Maybe
Yes
Yes
NA
NA
NA
microbial processes and chemical reactions can also be
reductive dechlorinations.
3.1.2 General Factors Influencing Subsurface
Biodegradation and NA
Subsurface microbes catalyze redox reactions in ground
water which alters the redox potential and impacts
the occurrence and rate of biotic transformations of
contaminants.
Under anaerobic environments, reducing compounds,
such as organic carbon, are fermented to produce
H2 which serves as e" donor for Dehalococcoides
and other dechlorinating bacteria (Duhamel et
al., 2002). Research has demonstrated that under
strongly reducing conditions in the presence of
sufficient supply of bioavailable natural organic
carbon, complete reductive dechlorination of PCE was
observed (Thomas et al., 2013). Therefore, dissolved
H2 concentrations could also be measured and used
to indicate the predominant microbially catalyzed
redox reactions and conditions in anoxic ground water
(Lovley et al., 1994). There may be competition for
H2 or other electron donor, or for electron acceptor,
between different microbial species carrying out
one or more of these processes, which can affect the
occurrence and extent of contaminant transformation
by a particular species.
The concentration of a target contaminant can also
impact the occurrence and rate of biodegradation. At
high enough concentrations, the contaminant may
be toxic to the microbes that degrade it, and low
concentrations may be insufficient to support growth
of the microbe. A co-existing contaminant may be
toxic or detrimental to a biodegradative process carried
out by specific bacteria.
At some sites, PCE, TCE, and TCA may be present as
a dense non-aqueous phase liquid (DNAPL) that acts
as a continual source of dissolved solvent as it dissolves
into the ground water. High dissolved concentrations
resulting from dissolution of the DNAPL contaminant
constituent may inhibit or prevent biodegradation.
The presence of a source, and especially DNAPL,
impacts the determination of attenuation rate
constants, and the source decay needs to be considered
(Newell et al., 2002).
3.2 PCE and TCE
3.2.1 Processes and pathways
Biotransformation of PCE and TCE is discussed
together, since they share many similar processes
(Table 2a and b).
The major biodegradation route of PCE and TCE is
through reductive dechlorination, a process known
as "halorespiration". During this growth-supporting
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 9
-------
microbial process, H2 is directly used as an e" donor
and the chlorinated solvent serve as the e" acceptor.
The H2 is produced during biodegradation of other
organic compounds, either naturally occurring organic
carbon or organic contaminants such as petroleum
hydrocarbons (Wiedemeier et al., 1999). PCE or TCE
loses a chlorine atom and is reduced. PCE and TCE
degradation products from reductive dechlorination
are DCE and the more toxic VC; however, the desired
end products are ethene, ethane, and ultimately CO2.
This biotransformation sequence may slow or stop at
DCE, with build-up of DCE concentrations (known
as "DCE stall"). In some cases, VC formed from DCE
may persist, if reducing conditions are not strong
enough. However, VC is biodegraded under aerobic
conditions more than the other chlorinated ethenes,
raising the possibility of its biodegradation as it moves
downgradient into a more aerobic environment. DCE
and VC can be biotically transformed through several
different mechanisms under either aerobic or anaerobic
conditions (Table 3).
Relevant coupled redox half reactions (modified from
Wiedemeier et al., 1998) for the PCE/TCE reductive
dechlorination sequence, and associated stoichiometric
concentration changes are:
PCE to TCE: CI2C=CCI2 + H+ + 2e = HCIC=CCI2 + Cl'
1 mg/L-> 0.79 mg/L
TCE to c-DCE: HCIC=CCI2 + H+ + 2e = HCIC=CCIH + Cf
1 mg/L-> 0.74 mg/L
c-DCE to VC: HCIC=CCIH + H+ + 2e = HCIC=CH2 + Cf
1 mg/L -> 0.64 mg/L
VC to ethene: HCIC=CH2 + H+ + 2e = H2C=CH2 + CI"
The predominant biotic transformation of the
parent compounds PCE and TCE that occurs and
that is desirable for remediation through NA is the
reductive dechlorination sequence PCE —» TCE —»
DCE —» VC —» non-toxic end products. However,
sufficient electron donors need to be present, along
with the requisite microbes. If not, the reductive
dechlorination sequence will be incomplete and result
in persistence of one or more of the contaminants.
3.2.2 Factors influencing transformation to
desired end product
The primary factors affecting the transformation of
PCE and TCE to innocuous end products (i.e., CO2
and Cl"1), and without accumulation of r-DCE and/or
VC, are (1) the presence of sufficient e" donor to drive
the redox conditions to the most efficient reductive
dechlorination processes, and (2) the presence of the
microbes necessary for the complete transformation.
The predominant redox condition affects the occur-
rence, type, and efficiency of the biotransformation
reaction which will occur for the chlorinated ethenes.
A highly reducing condition may be necessary for
efficient reductive dechlorination of VC to ethene.
Halorespiration is most efficient under sulfate-
reducing and methanogenesis, less efficient under
iron-reducing, and questionable under manganese-
reducing conditions (Bradley and Chapelle, 2010).
Halorespiration does not occur under aerobic or
nitrate-reducing conditions (North Wind, 2003), but
TCE reductive dechlorination to cis-DCE can occur
under iron-reducing conditions (Bradley and Chapelle,
2010). At contaminated sites where either geochemical
conditions are not appropriate for complete
anaerobic biodegradation of chlorinated ethenes or
Dehalococcoides ethenogenes microorganisms capable
of carrying out the transformation to ethene are not
present, direct aerobic biodegradation of VC offers a
remedial solution for persistent VC plumes that are
not amenable to the anaerobic process of reductive
dechlorination.
The final e" donor (H2) for the halorespiration process
to occur is produced through fermentation of organic
compounds. As discussed earlier, sufficient e" donors
must also be available for the redox conditions to reach
those in which reductive dechlorination occurs. At
many sites, the initial e" donor (from which the H2
ultimately comes from) is not identified, unless there
is a petroleum hydrocarbon (i.e., e" donor) plume
commingled with the chlorinated solvent plume.
Otherwise, the e" donor may be simply identified as
dissolved TOC.
The presence of the appropriate microbes, specifically
Dehalococcoides ethenogenes (DHC), is required for
10 Ground Water Issu
mthesis Report on State of Understanding of Chlorinated Solvent Transformation
-------
Table 3a. Compilation of compilations of chlorinated solvent biotic transformation first-order rate constants.
Contaminant: PCE
Type of
Study
Field
Field
Lab
Lab
and
Field
Field
Lab
Field
Lab
Field
Lab
and
Field
Field
Lab
and
Field
Field
Lab
Lab
and
Field
Biogeochemical
Conditions
Reductive
dechlorination
Not specified
Not specified
Anaerobic
Methanogenic
Methanogenic
Sulfate reducing
Sulfate reducing
Anaerobic
Anaerobic
Not specified
All studies
Aerobic
oxidation
Aerobic
oxidation
Aerobic
oxidation
First-Order Rate Constants (day"1)
Min
0.0022
0.0381
0
0.0007
0
0.0035
0
0.0002
0.0000
0
0
0
0
25th
0.0025
0.0005
0
0
Median
0.0030
0.003
0.00186
0.0007
0.0084
0.0041
0.0065
0.0006
0.009
0
75th
0.0047
0.0007
0.079
0.002
Max
0.0066
0.0381
0.071
0.034
0.071
0.0046
0.013
0.0029
0.0027
0.410
0
0.004
0.004
Mean
0.0038
0.0029
0.0265
0.0041
0.0204
0.0029
0.051
0
0.001
0.001
n, number
of studies
3
5
3
2
3
16
36
9
50
3
7
10
Reference
Aziz et al. 2000
Aziz et al. 2000
Aziz et al. 2000
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
Lawrence 2006
Lawrence 2006
Newell et al.
2006
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Notes
Table B-1. Used
Biochlor with
rates from AFCEE
database of 24
sites.
Fable 12. Median
of field values.
Cites Weidemeier
etal. 1999.
Fable 12. Range
of laboratory
values. Cites
Weidemeier et al.
1999.
Fable 2.1. Update
of Aronson and
Howard 1997.
Table E-12
Table D-12
Table E-12
Table D-12
Fable 15. Mean
of field/in situ
studies. Cites
Aronson and
Howard 1997.
Fable 15. Mean
or range for all
studies. Cites
Aronson and
Howard 1997.
Fable 8. Rate
constants are from
concentration vs.
time at a point.
Table 8
Table 7
Table 7
Table 8
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 11
-------
Table 3a. continued...
Lab
Field
Lab
Lab
and
Field
Lab
and/or
Field
Lab
and/or
Field
Lab
and/or
Field
Field
Aerobic
cometabolism
Reductive
dechlorination
Reductive
dechlorination
All studies
Reductive
dechlorination:
nitrate-reducing
Reductive
dechlorination:
iron-reducing
Reductive
dechlorination:
methanogenesis
Anaerobic
0
0
0
0
0
0.00019
0.002
0.013
0.004
0.080
0.050
0.147
0.054
0.080
0.410
1.96
0.410
0.0033
0.025
0.010
0.101
1.41
0
0.004
0.100
0.0029
3
13
23
61
3
2
22
16
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Weidemeier et
al. 1999
Table 7
Table 7
Table 7
Table 8
Table 8
Table 8
Table 8
Table 6-7. Mean
is from field/in
situ studies. Min
and max are
"recommended"
rate constants.
Cites Aronson and
Howard 1997.
Contaminant: TCE
Type of
Study
Field
Field
Lab
Lab
and
Field
Field
Lab
Field
Biogeochemical
Conditions
Reductive
dechlorination
Not specified
Not specified
Anaerobic
Methanogenic
Methanogenic
Sulfate reducing
First-Order Rate Constants (day"1)
Min
0.0008
0.0001
0.00082
0.0004
0.0020
0.0001
25th
0.0014
Median
0.0033
0.003
0.0016
0.0006
0.0145
0.0015
75th
0.0066
Max
0.0088
0.3452
0.04
0.0008
0.0400
0.0071
Mean
0.0041
0.0013
0.0170
0.0019
n, number
of studies
10
6
4
10
Reference
Aziz et al. 2000
Aziz et al. 2000
Aziz et al. 2000
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
Notes
Table B-1. Used
Biochlor with
rates fromAFCEE
database of 24
sites.
Table 12. Median
of field values.
Cites Weidemeier
etal. 1999.
Table 12. Range
of laboratory
values. Cites
Weidemeier et al.
1999.
Table 2.1. Update
of Aronson and
Howard 1997.
Table E-15
Table D-15
Table E-15
12 Ground Water Issu
mthesis Report on State of Understanding of Chlorinated Solvent Transformation
-------
Table 3a. continued...
Lab
Field
Lab
and
Field
Field
Lab
and
Field
Field
Lab
Lab
and
Field
Field
Lab
Lab
and
Field
Field
Field
Lab
Field
Lab
Lab
and/or
Field
Lab
and/or
Field
Lab
and/or
Field
Sulfate reducing
Anaerobic
Not specified
Not specified
All studies
Aerobic
oxidation
Aerobic
oxidation
Aerobic
oxidation
Aerobic
cometabolism
Aerobic
cometabolism
Aerobic
cometabolism
Aerobic/
Anaerobic
Reductive
dechlorination
Reductive
dechlorination
Anaerobic
oxidation
Anaerobic
oxidation
Reductive
dechlorination:
iron-reducing
Reductive
dechlorination:
sulfate-reducing
Reductive
dechlorination:
methanogenesis
0
-0.0010
0
0
0
0.105
0.024
0.024
0
0
0
0.002
0
-0.0001
0
0.2
0.001
0.005
0.001
0.0029
0.0003
0
0.26
0.002
0.008
0.004
0.0007
0.003
0.88
0.004
0.018
0.008
0.0110
0.0016
3.130
0.028
0.028
1.410
1.650
1.650
0.023
3.130
0.011
0.023
0.109
0.0049
0.0025
0.0006
0.173
0.006
0.005
0.948
0.509
0.586
0.003
0.196
0.003
0.011
0.015
7
30
78
13
86
2
10
11
3
14
17
1
32
24
11
7
10
HydroGeoLogic,
Inc. 1999
Lawrence 2006
Lawrence 2006
Newell et al.
2006
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Table D-15
Fable 15. Mean
of field/in situ
studies. Cites
Aronson and
Howard 1997.
Fable 15. Mean
or range for all
studies. Cites
Aronson and
Howard 1997.
Fable 8. Rate
constants are from
concentration vs.
time at a point.
Table 7
Table 7
Table 7
Table 8
Table 7
Table 7
Table 8
Table 7
Table 7
Table 7
Table 7
Table 7
Table 8
Table 8
Table 8
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 13
-------
Table 3a. continued...
Lab
and/or
Field
Field
Reductive
dechlorination:
mixed
Anaerobic
0.00014
0.0025
0.001
0.0025
2
47
Suarez and Rifai
1999
Weidemeier et
al. 1999
Table 8
Table 6-7. Mean
is from field/in
situ studies. Min
and max are
"recommended"
"ate constants.
Cites Aronson and
Howard 1997.
Contaminant: c;s-DCE
Type of
Study
Lab
Lab
and
Field
Field
Lab
Field
Field
Lab
Biogeochemical
Conditions
Not specified
All studies
Aerobic
cometabolism
Aerobic
cometabolism
Aerobic/
Anaerobic
Reductive
dechlorination
Reductive
dechlorination
First-Order Rate Constants (day"1)
Min
0.0086
0
0.281
0.081
0
0
0.001
25th
Median
75th
Max
0.0256
1.960
1.960
0.434
0.008
0.130
0.200
Mean
0.004
0.885
0.187
0
0.002
0.014
n, number
of studies
34
3
2
4
17
8
Reference
Aziz et al. 2000
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Notes
Fable 12. Range
of laboratory
values. Cites
Weidemeier et al.
1999.
Table 7
Table 7
Table 7
Table 7
Table 7
Table 7
Contaminant: DCE
Type of
Study
Field
Field
Biogeochemical
Conditions
Reductive
dechlorination
Not specified
First-Order Rate Constants (day"1)
Min
0.0003
0.0000
25th
0.0019
0.0000
Median
0.0033
0.0004
75th
0.0060
0.0000
Max
0.0573
0.0005
Mean
0.0096
n, number
of studies
9
2
Reference
Aziz et al. 2000
Newell et al.
2006
Notes
Table B-1 . Used
Biochlor with
"ates from AFCEE
database of 24
sites.
Fable 8. Rate
constants are from
concentration vs.
time at a point.
Contaminant: DCE (not cis)
Type of
Study
Lab
and
Field
Biogeochemical
Conditions
All studies
First-Order Rate Constants (day"1)
Min
0
25th
Median
75th
Max
1.150
Mean
0.149
n, number
of studies
27
Reference
Suarez and Rifai
1999
Notes
Table 7
14 Ground Water Issu
mthesis Report on State of Understanding of Chlorinated Solvent Transformation
-------
Table 3a. continued...
Field
Lab
Field
Field
Lab
Aerobic
cometabolism
Aerobic
cometabolism
Aerobic/
Anaerobic
Reductive
dechlorination
Reductive
dechlorination
0.390
0
0.001
0.010
1.150
0.714
0.006
0.270
0.720
0.196
0.003
0.101
4
4
16
3
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Table 7
Table 7
Table 7
Table 7
Table 7
Contaminant: DCE (all isomers)
Type of
Study
Lab
and
Field
Lab
and
Field
Lab
and
Field
Lab
and/or
Field
Lab
and
Field
Lab
and/or
Field
Biogeochemical
Conditions
All studies
Aerobic
cometabolism
Reductive
dechlorination:
iron-reducing
Reductive
dechlorination:
sulfate-reducing
Reductive
dechlorination:
methanogenesis
Reductive
dechlorination:
mixed
First-Order Rate Constants (day"1)
Min
0
0
0
0.002
25th
0.002
0.081
0.001
0.007
Median
0.004
0.434
0.002
0.016
75th
0.050
0.714
0.003
0.058
Max
1.96
1.96
0.005
0.200
Mean
1.41
0.591
0.002
0.045
0.047
0.001
n, number
of studies
61
13
8
3
8
2
Reference
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Notes
Table 8
Table 8
Table 8
Table 8
Table 8
Table 8
Contaminant: VC
Type of
Study
Field
Field
Lab
Lab
and
Field
Biogeochemical
Conditions
Reductive
dechlorination
Not specified
Not specified
Anaerobic
First-Order Rate Constants (day"1)
Min
0.0011
0.0003
0
25th
0.0016
Median
0.0047
0.0079
0.00405
75th
0.0134
Max
0.0334
0.01
0.0082
Mean
0.0099
n, number
of studies
7
Reference
Aziz et al. 2000
Aziz et al. 2000
Aziz et al. 2000
HydroGeoLogic,
Inc. 1999
Notes
Table B-1 . Used
Biochlor with
"ates from AFCEE
database of 24
sites.
Fable 12. Median
of field values.
Cites Weidemeier
etal. 1999.
Fable 12. Range
of laboratory
values. Cites
Weidemeier et al.
1999.
Fable 2.1. Update
of Aronson and
Howard 1997.
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 15
-------
Table 3a. continued...
Field
Lab
Field
Lab
Lab
and
Field
Field
Lab
Field
Lab
Lab
and
Field
Field
Field
Lab
Lab
Field
and
Lab
Field
Methanogenic
Methanogenic
Sulfate reducing
Sulfate reducing
All studies
Aerobic
oxidation
Aerobic
oxidation
Aerobic
cometabolism
Aerobic
cometabolism
Aerobic
cometabolism
Aerobic/
Anaerobic
Reductive
dechlorination
Reductive
dechlorination
Anaerobic
oxidation
Anaerobic
oxidation:
iron- reducing
Anaerobic
0.0005
0
0.0057
0
0.043
1.500
0.055
0.055
0.001
0
0
0.008
0.001
0.00033
0.005
0.064
0.576
0.008
0.002
0.0008
0.0076
0.051
0.091
1.500
0.012
0.163
0.114
1.960
0.073
0.006
0.0013
0.0082
8.020
0.125
1.960
0.576
8.020
0.009
0.007
0.520
0.120
0.120
0.0072
0.002
0.0008
0.0076
0.518
0.087
1.730
0.316
2.422
0.004
0.003
0.303
0.049
0.042
0.0079
2
0
1
2
27
4
2
2
5
3
4
4
6
7
19
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifia
1999
Suarez and Rifia
1999
Weidemeier et
al. 1999
Table E-16
Table D-16
Table E-16
Table D-16
Table 8
Table 7
Table 8
Table 7
Table 7
Table 8
Table 7
Table 7
Table 7
Table 7
Table 8
Table 6-7. Mean
is from field/in
situ studies. Min
and max are
"recommended"
"ate constants.
Cites Aronson and
Howard 1997.
Contaminant: TCA
Type of
Study
Field
Field
Biogeochemical
Conditions
Reductive
dechlorination
Not specified
First-Order Rate Constants (day"1)
Min
0.0044
25th
0.0055
Median
0.0066
0.0159
75th
0.0077
Max
0.0088
Mean
0.0066
n, number
of studies
2
Reference
Aziz et al. 2000
Aziz et al. 2000
Notes
Table B-1 . Used
Biochlor with
-atesfromAFCEE
database of 24
sites.
Fable 12. Median
of field values.
Cites Weidemeier
etal. 1999.
16 Ground Water Issu
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Table 3a. continued...
Lab
Lab
and
Field
Field
Lab
Field
Lab
Lab
Lab
Lab
Lab
Lab
Lab
Lab
and
Field
Field
Lab
Not specified
Anaerobic
Methanogenic
Methanogenic
Sulfate reducing
Sulfate reducing
Aerobic, 0.1
and 0.5 mg L"1
TCA
Nitrate-
reducing, 0.1
and 0.5 mg L"1
TCA
Sulfate
reducing, 0.1
mg L'1 TCA
Sulfate
reducing, 0.5
mg L1 TCA
Methanogenic,
0.1 mg L'"1 TCA
Methanogenic,
0.5 mg L""1 TCA
Anaerobic
Not specified
Sulfate reducing
0.0099
0
0
0.0034
0
0
0.00355
0.011
0.0065
0.0030
0.0092
0.0099
0.041
0.0
0.015
0.010
0.015
0.0182
0.0065
0.043
0.0064
No biotransformation observed.
No biotransformation observed.
0.239
0.0006
0.0003
0.0007
0.0009
0.0015
0.3013
0.0017
0.0162
0.0035
0.0142
0.0033
0.0013
5
2
3
3
1
1
1
1
1
1
28
6
1
Aziz et al. 2000
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
HydroGeoLogic,
Inc. 1999
Klecka et al.
1999
Klecka et al.
1999
Klecka et al.
1999
Klecka et al.
1999
Klecka et al.
1999
Klecka et al.
1999
Lawrence 2006
Newell et al.
2006
Scheutz et al.
2011
Fable 12. Range
of laboratory
values. Cites
Weidemeier et al.
1999.
Fable 2.1. Update
of Aronson and
Howard 1997.
Table E-13
Table D-13
Table E-13
Table D-13
Fable 2. Used
field soil and
ground water.
Fable 2. Used
field soil and
ground water.
Fable 2. Used
field soil and
ground water.
Dseudo-first-order
"ate constant.
Fable 2. Used
field soil and
ground water.
Dseudo-first-order
"ate constant.
Fable 2. Used
field soil and
ground water.
Dseudo-first-order
"ate constant.
Fable 2. Used
field soil and
ground water.
Dseudo-first-order
"ate constant.
Fable 15. Mean
or range for all
studies. Cites
Aronson and
Howard 1997.
Fable 8. Rate
constants are from
concentration vs.
time at a point.
Fable 3. Pseudo-
first-order
"ate constant.
1,1-DCAwas end
oroduct.
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 17
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Table 3a. continued...
Lab
Lab
and
Field
Field
Lab
Field
and
Lab
Field
Lab
Field
Field
Lab
Lab
and/or
Field
Lab
and/or
Field
Lab
and/or
Field
Field
Methanogenic
All studies
Aerobic
oxidation
Aerobic
oxidation
Aerobic
oxidation
Aerobic
cometabolism
Aerobic
cometabolism
Aerobic/
Anaerobic
Reductive
dechlorination
Reductive
dechlorination
Reductive
dechlorination:
nitrate-reducing
Reductive
dechlorination:
sulfate-reducing
Reductive
dechlorination:
methanogenesis
Anaerobic
0.0038
0
0
0
0
0
0
0
0.003
0.0013
0
0
0.002
0
0.025
0.010
0
0.013
0
0.125
0.195
0
0.038
0
0.880
0.0148
2.330
0.022
0.022
1.180
0.125
2.330
0
2.330
0.01
0.261
0.003
0.002
0.247
0.029
0.551
0
0.010
0.498
0.016
1
47
2
9
11
5
10
21
4
2
17
15
Scheutz et al.
2011
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Weidemeier et
al. 1999
Table 3. Pseudo-
first-order
rate constant.
1,1 -DCA was end
product.
Table 8
Table 7
Table 7
Table 8
Table 7
Table 8
Table 7
Table 7
Table 7
Table 8
Table 8
Table 8
Table 6-7. Mean
is from field/in
situ studies. Min
and max are
"recommended"
rate constants.
Cites Aronson and
Howard 1997.
Contaminant: 1,1 -DCA
Type of
Study
Field
Lab
Biogeochemical
Conditions
Reductive
dechlorination
Not specified
First-Order Rate Constants (day"1)
Min
0.0005
0.0044
25th
0.0005
Median
0.0008
75th
0.0019
Max
0.0033
0.0096
Mean
0.0014
n, number
of studies
3
Reference
Aziz et al. 2000
Aziz et al. 2000
Notes
Table B-1 . Used
Biochlor with
rates from AFCEE
database of 24
sites.
Table 12. Range
of laboratory
values. Cites
Weidemeier et al.
1999.
18 Ground Water Issu
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Table 3a. continued...
Contaminant: DCA (all isomers)
Type of
Study
Lab
and
Field
Lab
Lab
Field
Field
Lab
Field
Field
Biogeochemical
Conditions
All studies
Aerobic
oxidation
Aerobic
cometabolism
Aerobic/
Anaerobic
Reductive
dechlorination
Reductive
dechlorination
Reductive
dechlorination:
sulfate-reducing
Reductive
dechlorination:
methanogenesis
First-Order Rate Constants (day"1)
Min
0
0.014
0
0.028
0
25th
0
0.019
0
Median
0.001
0.047
0
75th
0.014
0.123
0.001
Max
0.131
0.131
0.011
0.044
0.028
Mean
0.017
0.067
0.002
0.036
0.003
0.006
n, number
of studies
25
2
5
16
2
13
3
Reference
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Suarez and Rifai
1999
Notes
Table 8
Table 7
Table 8
Table 7
Table 7
Table 7
Table 8
Table 8
Contaminant: CA
No experiments or results reported.
Contaminant: Dioxane
Type of
Study
Field
Lab
Lab
Biogeochemical
Conditions
Methanogenic,
<15°C, pH 6-8
Aerobic, 4 or
14°C, 50 mg L"1
dioxane
Aerobic, 14°C,
500 |ig L"1
dioxane,
CB1190
bacterial strain
First-Order Rate Constants (day"1)
Min
25th
Median
0
75th
Max
Mean
0
No significant dioxane biodegradation.
0.1
n, number
of studies
1
1
1
Reference
HydroGeoLogic,
Inc. 1999
Lietal. 2010
Lietal. 2010
Notes
Table E-39
Used microcosms
without
oioaugmentation
or substrate
addition, to
simulate natural
attenuation
conditions. High
concentration
simulated source
zone.
Used
microcosms with
bioaugmentation
and substrate
addition. Low
concentration
simulated leading
edge of plume.
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 19
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Table 3a. continued...
Lab
Aerobic, 14°C,
500 ug L"1
dioxane, DVS
5a1 bacterial
strain
0.4
1
Li etal. 2010
Used
microcosms with
bioaugmentation
and substrate
addition. Low
concentration
simulated leading
edge of plume.
Notes:
1 . Rows without rate constant data indicate biogeochemical conditions where no data was provided, and are left in for compari-
son to other conditions.
2. Description of biogeochemical conditions is as specific as was reported in the cited Reference.
20 Ground Water Issu
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Table 3b. Chlorinated solvent biotic transformation zero-order rates.
Contaminant: PCE
Type of
Study
Lab
Biogeochemical
Conditions
Reductive
dechlorination
Zero-Order Rate (ug L"1 day"1)
Min
13
25th
288
Median
577
75th
1040
Max
19800
Mean
1863
n, number
of studies
or rates
29
Reference
Suarez and Rifai
1999
Notes
Table 6
Contaminant: TCE
Type of
Study
Lab
Biogeochemical
Conditions
Reductive
dechlorination
Zero-Order Rate (ug L"1 day"1)
Min
314
25th
511
Median
760
75th
1297
Max
7490
Mean
1740
n, number
of studies
or rates
7
Reference
Suarez and Rifai
1999
Notes
Table 6
Contaminant: cis-DCE
Type of
Study
Lab
Biogeochemical
Conditions
Reductive
dechlorination
Zero-Order Rate (ug L"1 day"1)
Min
13
25th
183
Median
511
75th
1318
Max
16958
Mean
1854
n, number
of studies
or rates
18
Reference
Suarez and Rifai
1999
Notes
Table 6
Contaminant: DCE (not cis)
Type of
Study
Lab
Biogeochemical
Conditions
Reductive
dechlorination
Zero-Order Rate (ug L"1 day"1)
Min
9
25th
23
Median
250
75th
1385
Max
3470
Mean
850
n, number
of studies
or rates
8
Reference
Suarez and Rifai
1999
Notes
Table 6
Contaminant: VC
Type of
Study
Lab
Biogeochemical
Conditions
Reductive
dechlorination
Zero-Order Rate (ug L"1 day"1)
Min
2
25th
6
Median
11
75th
75
Max
495
Mean
107
n, number
of studies
or rates
9
Reference
Suarez and Rifai
1999
Notes
Table 6
Contaminant: Dioxane
Type of
Study
Lab
Biogeochemical
Conditions
Aerobic, 14°C,
500 ug L"1
dioxane
Zero-Order Rate (ug L"1 day"1)
Min
25th
Median
75th
Max
Mean
1.4
n, number
of studies
1
Reference
Li etal. 2010
Notes
Used microcosms
without
bioaugmentation
or substrate
addition, to
simulate natural
attenuation
conditions. Low
concentration
simulated leading
edge of plume.
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 21
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the complete reductive dechlorination sequence of
PCE to ethene (Figure A). A detailed discussion of the
role of DHC is provided by Maymo-Gatell (1997).
If DHC is not present, other microbes may partially
dechlorinate the PCE and/or TCE to DCE and/or
VC. If yet other appropriate microbes are present, the
products of the PCE and/or TCE dechlorination (i.e.,
r-DCE and/or VC) could be further dechlorinated.
Extreme pH or temperatures out of the range suitable
for efficient microbial activity may inhibit PCE and/or
TCE biotransformation. A pH range of between pH
5 and 9 has been cited (Wiedemeier et al., 1999) for
reductive transformation, as a screening measure.
Co-contaminants or interfering compounds may
have an inhibiting effect on biotransformation of a
target chlorinated solvent. These compounds may
include solvent stabilizers [e.g., up to about 5%
1,4-dioxane in TCA, or a large number of compounds
up to a total of about 1% in TCE (Mohr 2001)].
During biodegradation of PCE, Aulenta et al. (2006)
reported that the presence of a co-contaminant
1,1,2,2-tetrachloroethane (TCA) negatively impacted
the dechlorination of VC to ethene by DHC species.
Carbon tetrachloride (CT), but not the TCA, however,
inhibited PCE and VC biodegradation by the same
culture, even though it was able to cometabolize both
CT and TCA (literature cited by Aulenta et al., 2006).
TCA completely inhibited dechlorination of VC to
ethene in presence of a TCE-dechlorinating culture
as reported by Duhamel et al. (2002) and similarly,
the reductive dechlorination of PCE, TCE, cis-DCE,
and/or VC was partially or completely inhibited by
chloroform (CF) with a dechlorinating culture related
to DHC (Duhamel et al., 2002).
3.2.3 Geochemical conditions and
contaminant concentrations (required
measurements)
Geochemical conditions (e.g., redox conditions)
strongly influence which transformation processes
will occur and to what extent, as discussed above.
The naturally occurring e" acceptor(s) supply can
also impact the biotransformation process due to
competition with the chlorinated ethene e" acceptor.
The contaminant concentration may become
important in terms of microbial toxicity and e"
acceptor supply as mentioned in section 3.1.2.
Aulenta et al. (2006) identified some PCE- and TCE-
dehalorespiring bacterial strains that are inhibited by
PCE concentrations over 0.1 to 0.7 mmol L"1.
3.2.4 Indicator species - biological (required
measurements)
Dehalococcoides ethenogenes strain 195 (DHC) is
recognized (Maymo-Gatell, 1997; Maymo-Gatell et
al., 1999) as being capable of completely degrading
PCE to ethene, through the intermediate products
TCE, cis-DCE, trans-DCE, VC, and 1,1-DCE.
Other Dehalococcoides strains and known microbial
consortia (Wiedemeier et al., 1998; and Aulenta et al.,
2006) that are capable of biotransforming portions
of the chlorinated ethene degradation sequence
are identified in Figure 3.1. Mixed cultures that
can reductively dechlorinate DCE and VC are also
described by Bradley and Chapelle (2010). Molecular
biological tools (MBTs) are available to examine
the presence of degradative enzymes (tceA, vcrA,
and bvcA) (Figure 3.2). vcrA activity is required for
complete degradation of PCE to ethene through an
energy yielding pathway. A combination of tceA and
bvcA may lead to complete degradation; however,
through cometabolic reactions that are usually slower
than that observed for vcrA. Molecular biological
tools are described in depth in ITRC (2011).
3.2.5 Rates of transformation
The presentation, analysis, and use of
biotransformation rate data is complicated by the
manner in which these data are presented in the
literature, since the kinetics of chlorinated ethene
solvent biotransformation in field and laboratory
studies has been described using Monod kinetics,
Michaelis-Menten kinetics, zero-order rates, and by
first-order rate constants (Rifai et al., 2001; Aulenta
et al., 2006). Rate information resulting from
laboratory microbial degradation experiments may be
described using different parameters than the simple
first-order rates and rate constants that can be derived
from field measurements. Further, summaries of
kinetic parameter values from the literature often are
not accompanied by a full range of geochemical and
hydrogeological parameter values that could help
in understanding or modeling MNA at a field site.
22 Ground Water Issu
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Microbiology of Reductive
Dechlorination of Chloroethenes
Can accumulate if requisite
bacteria are not present
* H 2H HCI H V
Cl Cl 2H HCI Cl H 2H HCI H H 2H HCI H * H 2H HCI H H
C = C ' C = C ^' C = C ^~\ C = C /^' C = C
/ \ \/N/ /\ \/N
Cl Cl \ Cl Cl / Cl Cl \ H Cl H H
PCE
TCE
ci5-1.2-DCE
Dehalobacter
Dehalospirillum
Desulfitobacterium
Desulfuromonas
Dehalococcoides
vc
Ethene
some strains of
Dehalococcoides
Figure 3.1. Bacterial species involved in dechlorination processes.
Contaminant Degradation Pathway
The reaction is energy yielding, directly benefiting the bacteria
The reaction is cometabolic, not directly benefiting the bacteria
(Source: RITS, Fall 2007)
Figure 3.2. Enzymes involved in dechlorination processes.
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatic
Ground Water Issue 23
-------
Table 3a provides a sample of first-order rate constants
from a number of literature compilations, as well as
from some individual studies. Table 3b provides some
zero-order rates.
3.2.6 Case studies
TCE is found at 1034 National Priority List (NPL,
Superfund) Sites, while PCE is found at 938 NPL
sites (ATSDR, 2011). The great majority of these
sites where remedial activities have occurred have
used "active" remedial technologies, rather than
the "passive" MNA technology. Where MNA has
been used, it has almost always been a component
of the overall remedy, in combination with other
technologies used prior to MNA or for other portions
of the site. It is difficult to identify many sites where
MNA has been the sole remedial technology (although
the 23 PCE, TCE, or TCA sites evaluated by Newell
et al. (2006), were reported to have not had any other
remediation or source depletion activities). For many
sites, this means that the NA processes may have
been likely to be impacted by the other activities.
Nonetheless, there have been numerous sites where
successful, comprehensive MNA studies have been
conducted and extensive information obtained on the
NA processes and rates (whether or not MNA was
ultimately selected and was successful as a remedy).
The discussions throughout this document have
alluded to MNA sites; the references cited can be
referred to for further information on these studies.
A sampling of sites includes the Twin Cities Army
Ammunition Plant Superfund Site, MN for TCE
and TCA; Air Force Plant 44, Tucson International
Airport Area Superfund Site, Tucson, AZ for TCE
and TCA; Picatinny Arsenal, NJ for TCE; Altus AFB,
Alms, OK for TCE; Plattsburgh AFB, Plattsburgh, NY
for TCE; Dover AFB Superfund Site, Area 6, Dover,
DE; Lakehurst NAES Superfund Site, Lakehurst,
NJ; Moffett Field Superfund Site, CA; St. Joseph
Superfund Site, MI; and England AFB, LA.
3.3 TCA
3.3.1 Processes and pathways
Biotic transformation of TCA has many similarities
with biotic transformation of PCE and TCE. This
section focuses on significant differences in processes
for TCA.
The biotransformation processes for TCA are shown
in Table 3, with the most significant process being
reductive dechlorination via growth-supporting
halorespiration (i.e., with TCA as the electron
acceptor) (Scheutz et al, 2011). Under anaerobic
methanogenic conditions, TCA is reductively
dechlorinated (relatively faster) to 1,1-DCA, which is
then dechlorinated (relatively slower) to chloroethane
(CA). Either of the two degradation products can
be the end product, depending on the subsurface
microbiological and/or geochemical conditions,
although CA has been observed to be the most
common end product (Scheutz et al., 2011). Some
mineralization of each of these compounds may occur,
although it is likely to be minor (Vogel and McCarty,
1987; Scheutz et al., 2011).
Aerobic and anaerobic cometabolic dechlorination of
TCA and 1,1-DCA can occur, but are not significant
in terms of using MNA as a remedy for TCA (Scheutz
et al, 2011). Direct aerobic oxidation of CA (but not
TCA or 1,1-DCA) has been reported (Scheutz et al.,
2011).
3.3.2 Factors influencing transformation to
desired end product
The factors affecting the transformation of TCA to
innocuous end products (i.e., ethene, or ultimately
to CO2 and Cl"), without accumulation of CA are
somewhat different than the factors that impact
PCE and TCE degradation to those end products.
While the presence of sufficient e" donor to drive
the subsurface to methanogenic conditions and the
appropriate microbes are necessary for transformation
of the individual contaminant, there is not one sole
set of conditions where complete dechlorination of
TCA to innocuous end products occurs (as with
methanogenic conditions and the presence of DHC
for PCE dechlorination). TCA to CA dechlorination
will occur under one set of conditions (methanogenic
with the presence of the appropriate Dehalobacter
bacteria), while CA will be degraded to the desired end
products under a different set of conditions, through
aerobic oxidation in the presence of sufficient dissolved
oxygen and the appropriate aerobic microbes (Scheutz
et al., 2011). Known microbial cultures are unable to
completely dechlorinate TCA to ethane (Scheutz et al.,
2011).
24 Ground Water Issu
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The presence of CT and TCA inhibited the
biotransformation of each other under anaerobic
methanogenic conditions (Adamson and Parkin,
1999).
3.3.3 Geochemical conditions and
contaminant concentrations (required
measurements)
As with PCE and TCE, the redox conditions will
strongly influence what transformation processes
will occur and their extent. Since the TCA reductive
dechlorination product CA can be aerobically
oxidized, identification of downgradient zones of
sufficient dissolved oxygen, and the evaluation of the
migration pathway of the CA, will be important to
help assure that this degradation product does not
persist.
3.3.4 Indicator species - biological (required
measurements)
As with PCE and TCE, the presence of suitable
microbes with the ability to transform TCA is
necessary, specifically, the appropriate Dehalobacter
species for the reductive dechlorination of TCA to 1,1-
DCA and/or CA.
3.3.5 Rates of transformation
TCA halorespiration (and anaerobic cometabolic
transformation) has been described using pseudo-
first-order kinetics (Scheutz et al., 2011), and with
Michaelis-Menten model parameters and first-order
rate constants (Rifai et al., 2001). Table 3a provides a
sample of first-order rate constants from a number of
literature compilations, as well as from some individual
studies. Table 3b provides some zero-order rates.
3.3.6 Case studies
TCA is found at 791 NPL Sites (ATSDR, 2011).
Many of these sites that have TCA also have PCE and/
or TCE. One well-studied TCA site is the Twin Cities
Army Ammunition Plant, MN (e.g., Wilson 2010).
Scheutz et al. (2011) discuss TCA biotransformation
under enhanced reductive dechlorination (ERD) at
18 sites where both TCA and chloroethenes were
found, and four sites with just TCA. Although these
sites used the active remedial technology of ERD, and
not MNA, baseline data was collected prior to ERD
implementation and indicated the potential for some
anaerobic dechlorination of the TCA via NA. Further
information on these, and other sites, can be found
in the cited references and may be available in a web
appendix to this document.
3.4 Dioxane
3.4.1 Processes and pathways
Dioxane biodegradation occurs through oxidation,
under aerobic conditions, in both growth-supporting
(i.e., as primary substrate) and non-growth-supporting
(i.e., cometabolic) processes involving certain
monooxygenase enzymes. Three bacterial strains and
one fungus have been identified that use dioxane for
growth, while a larger number of bacteria and one
fungus have been reported to degrade dioxane in the
presence of an alternate substrate (i.e., non-growth-
supporting; cometabolic) such as methane. The
dioxane degradation pathway proceeds to complete
mineralization. The initial degradation step is rate-
limiting, with subsequent degradation steps being
fast. Intermediate degradation products have been
identified (including ethylene glycol); however, these
products are further degraded and mineralization
ultimately occurs (Mahendra and Alvarez-Cohen,
2006; Mahendra et al., 2007; Mora and Chiang,
2011). This suggests that undesirable degradation
products will not occur or persist.
Dioxane biodegradation in laboratory experiments was
described by Monod kinetics (Mahendra and Alvarez-
Cohen, 2006) and by either zero-order kinetics for
"natural attenuation" treatment or first order for
bioaugmented treatments (Li et al., 2010), as shown in
Table 3a and b..
3.4.2 Factors influencing transformation to
desired end product
A variety of bacterial strains were able to use potential
co-contaminants as growth substrates for the
cometabolism of dioxane under laboratory conditions,
including toluene, tetrahydrofuran (THF), MTBE,
and methane (Mahendra and Alvarez-Cohen, 2006).
Although acetylene inhibited biodegradation of
dioxane as a growth substrate, after its removal and
when an alternate substrate was supplied, the ability
to biodegrade dioxane was restored (Mahendra and
Alvarez-Cohen, 2006).
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 25
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Li et al. (2010) simulated natural attenuation
conditions in laboratory microcosms to investigate
dioxane biodegradation at temperatures (4 and 14 °C)
that are lower than the typical laboratory conditions
(>20 °C). They also studied the impact of dioxane
concentration, using higher concentrations (50 mg L"1)
to represent a source zone, and lower concentrations
(500 iig L"1) to represent the leading edge of a
plume. No significant biodegradation occurred at
either temperature with the higher 50 mg L"1 dioxane
concentration. However, at the lower 500 tig L"1
dioxane concentration, significant biodegradation
was observed, with dioxane decreasing from 500 to
130 iig L"1 in six months.
3.4.3 Geochemical conditions and
contaminant concentrations - required
measurements
Dioxane biodegradation occurs under aerobic
conditions, requiring the presence of molecular oxygen
(Mahendra and Alvarez-Cohen, 2006), although
Mohr et al. (2010) cited laboratory studies in one
investigation that anaerobic biodegradation of dioxane
occurred under iron-reducing conditions. Field
measurements and identification of the aerobic and
anaerobic zones at a site are likely to indicate where
dioxane biodegradation is possible.
A wider variety of microbes are capable of dioxane
cometabolism than use it as a primary growth substrate
(Mahendra and Alvarez-Cohen, 2006). Therefore,
the identification of a primary substrate source, such
as methane, THF, or other cyclic ethers (Mora and
Chiang, 2011) should provide additional supporting
evidence for the potential occurrence of NA via biotic
transformation.
For all the compounds discussed in this document,
the contaminant concentration at a number of
longitudinal locations in a plume is an obvious
measurement, not only for calculation of attenuation
rates, but also in terms of potential toxicity issues.
There appears to be very limited literature on microbial
toxicity due to high dioxane concentrations; however,
Li et al. (2010) reported that significant dioxane
biodegradation occurred at 500 tig L"1, but not at
50 mg L"1. The number of measurement locations is
site-specific.
3.4.4 Indicator species - required biological
measurements
Research has suggested that dioxane is biodegraded by
Pseudonocardia dioxanivoram CB1190, Pseudonocardia
benzenivoram B5, and Rhodococcm strain 219 as a
sole source of carbon and energy (Mahendra and
Alvarez-Cohen, 2006). As apparent, only a limited
number of microbes are capable of utilizing dioxane
as a growth substrate. Thus, the identification of the
known dioxane-degrading microbes and even more
significantly, confirmation of their monooxygenase
enzymatic activity is the most important evidence for
potential MNA at a specific site. As indicated earlier,
dioxane can also be cometabolized by a larger number
of bacterial species, so the identification of those
bacteria could be advantageous.
3.4.5 Rates of transformation
There has been very little investigation, determination,
or reporting of dioxane biotic transformation rates
under field conditions. Mohr et al. (2010) summarize
laboratory research and present Monod kinetic
parameter values for 1,4-dioxane biodegradation,
including those in Mahendra and Alvarez-Cohen
(2006).
A zero-order rate of 1.4 + 0.02 tig L"1 day"1 was
calculated for biodegradation of 500 tig L"1 dioxane at
14 °C in laboratory microcosms containing soil and
ground water from a dioxane-contaminated site, under
simulated natural attenuation conditions (i.e., no
biostimulation or bioaugmentation) (Li et al., 2010).
3.4.6 Case studies
Dioxane has not been the primary or sole target for
MNA at contaminated sites, and has seldom been
included in the evaluation of MNA at chlorinated
solvent sites. The limited literature on dioxane and
MNA at contaminated sites is summarized below.
3.4.6.1. Mohr et al. (2010) presented seven
case studies of dioxane site investigations and
remediation. MNA does not appear to have been
considered for all or most of the sites. Each site had
some active remedial technology implemented. The
off-site plume beyond a ground-water extraction
system at one site may have been considered for
MNA; however, it was believed that any NA would
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be via dispersion, diffusion, and dilution rather than
via dioxane transformation.
3.4.6.2. Biotic NA of a very large dioxane plume at
a site near Wilmington, NC was hypothesized by
Chiang et al. (2008). The dioxane plume was the
result of releases during chemical manufacturing
activities at the site; the dioxane was not associated
with chlorinated solvents. Chiang et al. (2008)
stated that the results from a calibrated model,
and dioxane concentration declines observed
during long-term monitoring, indicated that "the
rate of dioxane attenuation...cannot be explained
solely due to nonbiological and abiotic attenuation
mechanisms", which suggested that there were
oo
biological "degradation mechanisms that have limited
the migration rate and size of the plume". They also
stated that the "extent of negative ORP and ferrous
iron" correlated with the locations of reduction in
dioxane concentrations, "suggesting the potential for
biological attenuation". However, no direct evidence
such as intermediate products or presence of
dioxane-degrading microbes were observed. Jenkins
et al. (2009) discussed the weaknesses in Chiang
et al.'s (2008) use of limited data and modeling
as the primary evidence of in situ biodegradation,
concluding that this site was not likely to be a
good candidate for MNA. The inappropriate or
premature conclusion that NA via biodegradation
was occurring at this site indicates the need for a
robust site characterization and collection of the
appropriate parameters.
3.4.6.3. An investigation of NA was conducted for
a large dilute TCE and 1,4-dioxane plume at the
Air Force Plant 44, Tucson International Airport
Area Superfund Site, Tucson, AZ. Both TCE
and TCA had been used at the site; however, the
main contaminants were the TCE (maximum
of 520 iig L"1) and dioxane (maximum of
1,110 iig L"1), which was present due to the use of
TCA. The site ground water was aerobic, with low
TOC. The plumes appeared to be shrinking, and
MNA was considered for part of the site remedial
strategy (Mora and Chiang, 2011). Pump-and-treat,
with reinjection of the treated ground water around
the plume perimeter, was started in 1987 (Chiang
et al., 2012); thus, the site has an active remedial
strategy that complicates calculation of dioxane
attenuation rates due solely to NA.
Site ground-water samples were tested using stable
isotope probing (SIP) with 13C-dioxane baited
bio-traps. Phospholipid fatty acid analysis (PLFA)
indicated that 13C was incorporated into microbial
biomass, detection of 13C in CO2 indicated that
some dioxane was mineralized, and quantitative real
time polymerase chain reaction (qPCR) indicated
the presence of the necessary bacteria and enzymes.
Enzyme activity probe analysis indicated that the
necessary toluene oxygenase enzymes were present
and active. This was the first field study to directly
indicate the natural biodegradation of dioxane
in the context of subsurface NA; however, the
analyses used in the study were unable to address
the determination of attenuation rates (Mora and
Chiang, 201 1; Chiang et al., 2012).
4. ABIOTIC TRANSFORMATIONS
The following discussion provides an overview
of current understanding of the pathways and
geochemical conditions controlling the abiotic
transformation of the contaminants of interest: PCE,
TCE, and 1,1,1 -TCA. This discussion will not address
1 ,4-dioxane due to the fact that there is no evidence in
the literature indicating that it is susceptible to abiotic
degradation.
The transformation of chlorinated solvents in the
subsurface is inextricably linked to a set of complex
biological, chemical and geochemical processes.
Overall transformation rate constants for chlorinated
solvents represent a contribution from both abiotic
and biological processes:
~~ ^
abiotic
Each of the rate constants represents the relative
contribution of degradation processes, which
depending on the chlorinated solvent, can include
both abiotic reduction and hydrolysis.
^abiotic = ^red ~"~ ^hyd
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The relative contribution of these terms is dependent
on the inherent reactivity of the chlorinated solvent
and the geochemical conditions of the aquatic
ecosystem of interest. The inherent reactivity is a
reflection of the strength of C-C1 bonds and the
reactivity of the abiotic reductants and nucleophiles
present in the aquatic system of interest.
As discussed below, the relative contributions of
kred and khyd is dependent on the structure of the
chlorinated solvent. PCE and TCE, are susceptible to
only abiotic reduction (i.e., k^^ = kK^), where as the
abiotic degradation 1,1,1 -TCA is susceptible to both
abiotic reduction and hydrolysis. The rate constant for
hydrolysis is dependent on the neutral and acid- and
base-catalyzed processes described by:
£hyd = £base [substrate] [OH~] + £n [substrate] +
4dd [substrate] [H+]
Consequently, based on values for the individual rate
constants, which can be measured relatively easily
in the laboratory, and the pH of the reaction system
of interest, it is a fairly straight forward process to
calculate the overall hydrolysis rate constant, £hyd.
The situation for predicting rate constants for abiotic
reduction is much more challenging, primarily
due to the fact that numerous abiotic reductants
may contribute to the overall rate constant, kred,
the formation and reactivity of which will vary as a
function of geochemical conditions. Furthermore, our
knowledge base at this point is dependent primarily
on laboratory studies of abiotic model systems and
anaerobic microcosms designed to mimic naturally
occurring conditions in the subsurface. The extent to
which these results studies apply to natural systems not
fully understood at this time.
The formation of abiotic reductants in anaerobic
aquifer systems is the result of the biologically-
mediated oxidation of bioavailable organic matter
resulting in the reduction of various e" acceptors (e.g.,
Fe(III) oxides and sulfate) as described by Terminal
Electron Accepting Processes (TEAPs). The resulting
redox zones in anaerobic subsurfaces can be mapped
by the measurement of solution phase species (e.g.,
Mn2+, Fe2+, H2S and CH4) resulting from reduction
of the e" acceptors (Bjerg, Rugge et al. 1995; Chapelle,
McMahon et al. 1995; Jeong and Hayes 2007;
Himmelheber, Taillefert et al. 2008; Himmelheber,
Thomas et al. 2008). Additional information
concerning the determination of redox zones in
anaerobic aquifers is provided from the measurement
of dissolved H2 concentrations based on a gas stripping
procedure (Lovley, Chapelle et al. 1994). EachTEAP
has a H2-utilizing efficiency resulting in characteristic
concentrations of dissolved H2 (i.e., < 0.1 nM H2 for
Nitrate reducing zones; 0.2 to 0.8 nM H2 for iron
reducing zones; Ito 4 nM H2 sulfate reducing zones,
and 5 to 15 nM H2 methanogenic (Lovley and
Goodwin 1988).
This concept provides a useful construct for the
subsequent discussion of the formation of abiotic
reductants in anaerobic subsurface systems. Although
laboratory studies can be designed to mimic specific
redox zones, their occurrence in natural systems is
often complex with overlapping and/or completely
mixed redox zones. One result of this scenario is the
difficulty in identifying the predominant chemical
reductants in these complex systems.
These reactive forms are primarily reactive surfaces
such as surface complexed Fe(II) and reactive minerals
such as green rusts and iron sulfides, all of which form
as the result of reactions of high concentrations of
ferrous iron and sulfide. These abiotic reductants that
are known to form as a function of iron and sulfate
redox zones are illustrated in Figure 4.1.
Nitrate
Reducing
C HCOj
X
NOj N2
Manganese
Reducing
Corg HC°3
X
MnO; Mn2*
Iron
Reducing
Corg HC°3
>-<^
Fe(OH)3 Fe2*
Sulfate
Reducing
C HC°3
°org
X
SO;' HS"
Methanogenic
Corg HC°3
V_>
CH4
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Abiotic Reductants
Surface Complexation Mineral Formation
Iron Bearing Hydroxides (Green Rusts):
GR-CI: [Fe2+Fe3+(OH)8]+ [CI»/?H2O]-
GR-SO4:[Fe2+Fe3+(OH)12]2[SO4»nH2O]-
GR-CO3: [Fe2+Fe3+(OH)12]2+ [CO3«nH2O]
I
O
I
Fe3+
Goethite
Surface
Iron Bearing Oxides:
Magnetite: Fe2+ + 2Fe3+ + 8OH' ^ Fe3O4
Iron Bearing Sulfides:
Mackinawite: Fe2+ + HS-
Pyrite: FeS + S° «-» FeS2
FeS + H+
Figure 4.1. Formation of abiotic reductants as a function of iron and sulphate reducing zones.
4.1 PCEandTCE
4.1.1 Processes and Pathways
The abiotic reduction of chlorinated solvents has
received much attention over the past decade due to
the observations that:
• Abiotic reduction pathways result in reaction prod-
ucts that are of much less concern than those based
on biologically-mediated reductive transformations
• The toxic reaction products formed from the
biologically-mediated process are susceptible to
abiotic degradation
• Lower concentrations of the targeted chlorinated
solvents can be achieved in remediation scenarios
by maximizing geochemical conditions for abiotic
degradation
The abiotic reductions of PCE andTCE have been
demonstrated in a number of abiotic model systems
and anaerobic microcosms designed to mimic iron-
reducing and sulfate-reducing zones in anaerobic
systems. Figure 4.2 illustrates the pathways for both
the abiotic and biologically-mediated reduction
of PCE and TCE. The abiotic pathway occurs
predominantly through reductive elimination
resulting in the formation of the reactive intermediate
dichloroaceteylene (Lee and Batchelor 2002).
Subsequent hydrogenolysis of dichloroacetylene
results in the formation of acetylene through the
reactive intermediate, chloroacetylene, which is
reduced further to ethane and ethene, all of which are
relatively innocuous degradation products (Butler and
Hayes 2001; Lee and Batchelor 2002). In contrast,
the biologically-mediated pathway is dominated by
hydrogenolysis (i.e., the replacement of a Cl group
with an H) to form TCE. Sequential hydrogenolysis
of TCE gives oy-l,2-DCE and subsequently VC, both
of which are susceptible to abiotic reduction resulting
in the formation of acetylene, ethene, and ethane.
4.1.2 Factors influencing transformation to
desired end product
The desired end products for the reductive
transformation of PCE andTCE are those for
which all of the Cl groups have been removed (i.e.,
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatic
jnd Water Issue 29
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Cl
Reductive
Elimination
cu
Dichloroacetylene
HC
Chloroacetylene
A
etheylene
Hydrogenolysis
H
JCE
H3C—CH3
ethane
Figure 4.2. Reaction Scheme illustrating the degradation pathways for PCE in anaerobic systems and the predomi-
nant processes controlling each of the transformation steps: A = abiotic degradation pathway, B = biotic
degradation pathway.
acetylene, ethene, and ethane). Consequently,
conditions that maximize the potential for abiotic
reduction, as discussed below, will favor the formation
of these desired end products. pH is also a factor
in determining the formation of the desired end
products as higher values (>8) increase rates of abiotic
transformations (see Table 4.1) and is thought to
inhibit the growth of dechlorinating bacteria.
4.1.3 Geochemical conditions
Our understanding of the geochemical conditions
controlling the abiotic reduction of PCE andTCE
is the result of laboratory based studies of abiotic
model systems and anaerobic microcosms. In total
these studies indicate that subsurface conditions
defined as iron and sulfate reducing will promote
abiotic reduction of PCE andTCE as a result of
30 Ground Wah
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Table 4.1. Surface area-normalized rate constants, /csa, with units of Lm~2day~1, for PCE, TCE, c/s-DCE and VC mea-
sured in anoxic model studies and anaerobic microcosms.
Exp#
1
2
3
4
5
6
7
8
9
10
12
13
14
15
Reaction System
FeS,
pH 7
FeS,
pH 8
FeS,
pH 9
GR-CI,
pH 8
FeS2,
pH 8
GR-SO4,
pH 8
Fe304,
pH 8
Fe(ll)/goethite,
pH 8
Microcosm,
pH 7
Microcosm,
pH 8
FeS2
Fe304
FeS,
pH=8.3
FeS,
pH=8.3
(0.04 M FeCI2)
PCE
(6.3± 1.6)x10'5
(5.3±0.5)x10'4
(1.21 ±0.1)x10'3
(5.6 ± 1.4)x10'6
(1.6± 1.0) x10'6
NC
NC
NC
(1.8± 1.2)x10'4
(9.1 ± 1.6)x10'4
2.0 x10'5
8.4 x10'7
(7.6 ± 1.0) x10'4
(3.8± 0.3) x 10"3
TCE
(1.6± 0.2) x10'4
(6.4 ± 0.8) x10'4
(2.9±0.61)x1Q-5
(6.4 ± 1.5)x10'5
NC
NC
NC
(6.2± 5.7)x10'4
(1.7± 1.9)x10'3
2.5 x10'5
7.2X10'7
(2.1 ± 0.1) x 10-3
(2.0± 0.1) x10"2
cis-DCE
NR
NR
NR
NR
NR
NR
NR
NR
NR
NR
1.3x10-5
5.6 x10'7
NR
NR
VC
NR
NR
NR
NR
NR
NR
NR
NR
NR
NR
2.27x10'5
5.6 x10'7
NR
NR
Reference
Butler,
Elizabeth
2009
Butler,
Elizabeth
2009
Butler,
Elizabeth
2009
Butler,
Elizabeth
2009
Butler,
Elizabeth
2009
Butler,
Elizabeth
2009
Butler,
Elizabeth
2009
Butler,
Elizabeth
2009
Butler,
Elizabeth
2009
Butler,
Elizabeth
2009
Leite, 2002
Lee, 2002
Jeong, 2007
Jeong, 2007
the formation of high concentrations of Fe(II) and
(S-II), and the subsequent formation of reactive
minerals as illustrated in Figure 4.1. Under most field
conditions, both abiotic and biologically-mediated
reduction of PCE and TCE will occur. The relative
rates of these processes will depend on the abundance
of dechlorination bacteria and the mass loadings of
reactive minerals.
4.1.4 Indicator species (chemical)
Indicator species (i.e., indicators of reactivity)
for abiotic reductions are those that will reflect
the reactivity of various abiotic reductants due to
formation and subsequent reactions of ferrous iron
and sulfide. This can include direct measures of the
reactive species (e.g., mass loadings of iron sulfides)
or measures of species that are not reactive, but
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 31
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may reflect the reactivity of the abiotic reductants
(e.g., aqueous phase concentration of ferrous iron
and sulfide). The direct measure of solid-phase
reactive species is challenging at best and is based
primarily on laboratory methods involving sequential
extraction methods. These extraction methods
will provide measures of weakly bound Fe(II) (i.e.,
surface complexed Fe(II) and strongly bound Fe(II),
acid-soluble sulfur, and chromium-extractable sulfur
(CrES), which provide measures of reactive Fe(II) and
S(-II) bearing minerals (Kostka and Luther III, 1994;
Heron, Bjerg et al. 1995).
By comparison, the measurement of aqueous phase
ferrous iron, sulfide and hydrogen concentrations are
quite feasible. Additionally, these parameters have
been measured for the principal aquifers in the U.S.
and are available in the USGS National Water Quality
Data Base. Although studies to determine the efficacy
of soluble ferrous iron and sulfide as indicators of
reactivity for abiotic reductive dehalogenation have
not yet been reported, aqueous phase concentrations
of ferrous iron measured in iron-reducing sediments
were shown to correlated strongly with the rates for the
abiotic reduction measured for a nitro aromatic probe
chemical in 21 iron-reducing sediments collected from
a diverse set of sites across the country.
When given enough time for reactions to proceed
to their maximum extent, reductive capacities are
defined by the amount of oxidant reduced. For PCE
reductions by active mineral reactions, reductive
capacities were found to correlate with the Fe(ll)
content (Lee and Batchelor 2003).
Also, an increase in reduction rate constants for PCE
and TCE in FeS systems treated with increasing
concentrations of Fe(II) has been reported, which was
attributed to an increase in the presence of different
types of solid-bound Fe phases with Fe(II).
4.1.5 Rates of transformation
Rate constants for abiotic reduction of PCE and
TCE have been measured in laboratory based abiotic
model systems and anaerobic microcosms designed
to mimic iron and sulfate reducing zones in natural
subsurface conditions. Abiotic degradation rate
constants for PCE and TCE measured in situ have not
been reported. A summary of pseudo-first-order rate
constants generated from these studies are summarized
in Table 4.1. These data are grouped according to
the study in which they were generated. Comparison
of rate constants generated from different studies is
somewhat problematic primarily due to the differences
in procedures used to generate the reactive minerals
resulting in materials with varying reactivity. Analysis
of the rate constants measured within a given study
does allow for a number of general observations as
reported below.
4.1.6 Normalization of rate constants to account
for partitioning
Rate constants for the abiotic transformation of PCE
and TCE in Table 4.1 were normalized for the effects
of partitioning among the gas, aqueous, and solid
phases according to:
k =_j«.
m,corr j-^
r
where F{, the partitioning factor, is defined as:
Kis is calculated as follows:
-K *
-AM7T~
where Ki>d is the solid/water distribution coefficient,
and^c is the weight fraction of organic matter in the
solid phase. Ki>d can be estimated from the empirical
relationship Ki4 = KitOCfoc.
These rate constants, which have been adjusted for
partitioning, were subsequently normalized to the
surface areas of the various reactive mineral phases,
providing surface normalized rate constants, £sa>
with units of L m~2day~1. The normalization of rate
constants for partitioning and surface area allow for
the direct comparison of reduction rates for PCE and
TCE measured in the various experimental systems.
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General observations based on the kinetic data
reported in Table 4.1:
• Rates for the abiotic reduction of PCE and TCE
increase with pH. PCE reduction by FeS increased
by approximately an order of magnitude with each
pH unit (Exps. 1, 2 and 3). Similar results were
observed for the microcosm studies (Exps. 9 and 10).
• Based on results of model studies of individual
reactive minerals (Exps, 2, 4, 5, 6, 7 and 8), their
relative reactivities can be assigned as follows:
FeS > GR-C1 ~ FeS2 > GR-SO4 ~ Fe3O4
The contribution of any one of these reductants
to the rate of abiotic reduction will depend on the
concentration and surface area of the reductant.
• The pathways for abiotic reduction as illustrated in
Figure 4.2 follow the same pathway (i.e., reductive
elimination) regardless of the relative contributions
of these abiotic reductants,
• Half lives for biodegradation (-10 days, not shown)
of PCE andTCE in the anaerobic microcosms were
shorter than those measured in abiotic systems of
reactive minerals, 900 to 5,000 days for PCE and
500-1,000 days for TCE. Tne half lives for abiotic
reduction were calculated from rate constants that
were mass normalized to FeS surface areas (Exps
9 and 10).
• Abiotic degradation, though slower than biodegra-
dation rates, can be significant when biodegradation
is not complete leading to the formation of cis-DCE
and VC (Exps 12 and 13).
Extrapolation of laboratory based generated rate constants
to field conditions
Lee and Batchelor have proposed a method for
extrapolating first-order rate constants measured in
model systems of reactive iron sulfides to aquifers
containing the reactive iron sulfide (Lee and Batchelor
2002). The following example is based on the
reduction kinetics measured for TCE in a suspension
of GR-SO4. A number of assumptions are required
for this extrapolation:
• The initial reductive capacity concentration (C°RC)
in the aquifer can be calculated by assuming that
green rusts represents 1% of the iron content of
the soil
• Based on the assumption that iron content is 2.6%,
a bulk density 1.4 kg/L, and a porosity of 0.40, the
mass of iron per volume water can be calculated
as 91 g/L
• Assuming that GR-SO4 is 52.6% iron, the green
rust concentration can be calculated as 1.73 g/L
• Based on measured rate constants in the GR-SO4
model systems, the calculated value for C°RC is
0.0225 mM
• Assuming a soil organic fraction of 0.005 and an
organic carbon partition coefficient of 206 L/kg,
a partition coefficient of 4.66 can be calculated
for TCE
Based on these assumptions, the following equation
was then used to calculate a pseudo-first-order rate
constant of 0.0037 day-1, which gives an apparent
half-life of 190 days in the simulated aquifer for the
reduction of TCE by GR-SO4:
*! =
(k IPCE (CRC j
l/K+C
RC
Where k is the experimentally determined pseudo-first-
order rate constant measured in the GR-SO4 model
system, PCE is the partition coefficient for partitioning
to the gas, aqueous and solid phases, Kis the sorption
coefficient, and C°RC is the initial reductive capacity
concentration.
This same approach was used to extrapolate results
from laboratory studies of PCE in suspensions of
pyrite and magnetite to estimate half lives for PCE of
13 days by pyrite and 608 days by magnetite under
field conditions (Lee and Batchelor 2003). These
results suggest that pyrite formed under sulfate
reducing conditions has the potential to significantly
contribute to the abiotic reduction of PCE.
4.2 TCA
4.2.1 Processes and Pathways
Relative to PCE andTCE, studies of the abiotic
degradation of TCA are limited. Figure 4.3 illustrates
the pathways for both the abiotic hydrolysis and
reduction, and biologically-mediated reduction of
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TCA based on our knowledge of the existing process
science (Vogel and McCarty 1987; Haag and Mill
1988; Butler and Hayes 2000; Gander, Parkin et al.
2002). In the case of TCA, base-catalyzed hydrolysis
results in the formation of 1,1-DCE through
elimination and acetic acid through nucleophilic
substitution(Haag and Mill 1988). Hydrogenolysis
mediated by both abiotic and biologically mediated
processes results in the formation of 1,1-DCA, and
subsequently CA, which is susceptible to hydrolysis to
form ethanol. Laboratory studies have demonstrated
that the formation of acetic acid occurs at a rate
- 5 times faster than the formation of 1,1 -DCE
(Haag, Mill et al. 1986). Although product recoveries
are typically quite low (< 10%) for the formation
of 1,1-DCA in FeS suspensions, 1,1-DCA was the
only product observed (Butler and Hayes 2000;
Gander, Parkin et al. 2002). With the addition of
a methanogenic consortium to the FeS suspensions,
product recovery of 1,1-DCA increased to -46%
(Gander, Parkin et al. 2002).
4.2.2 Factors influencing transformation to
desired end product
Of the three transformation pathways for TCA
illustrated in Figure 4.3, it is abiotic hydrolysis that
results in formation of the degradation product (i.e.,
acetic acid) of least concern. Because the hydrolysis
of TCA is base catalyzed, increases in pH will increase
the rate of TCA hydrolysis; however the rate of
elimination, which leads to the formation of 1,1-TCE,
will also increase with pH.
4.2.3. Geochemical conditions
The abiotic reduction of TCA in mackinawite (FeS)
suspensions suggest that sulfate-reducing conditions
will favor the abiotic reduction of TCA (Butler and
Hayes 2000; Gander, Parkin et al. 2002).
4.2.4 Rates of transformation
Reductive
Elimination
Cl
og
F
*
Hydropenolysis
A, B
Hydrolysis
0
1,1-DCE
H
-CH3
Acetic Acid
1,1-DCA
Hydrolysis
CA
Ethanol
Figure 4.3. Reaction scheme illustrating the degradation pathways for TCA in anaerobic systems and the predomi-
nant processes controlling each of the transformation steps: A = abiotic degradation pathway, B = biotic
degradation pathway.
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Based on the limited process science available, the
abiotic reduction of TCA is controlled primarily by
the presence of FeS in aquifer systems. The overall rate
term is characterized by a second-order rate term:
where {FeS} is the surface area concentration given by
the product of the mass concentration (S, g L"1). The
second-order rate constant, £FeS (L m"2d"1) is defined
as:
The £FeS determined by Gander, et. al. (Gander, Parkin
et al. 2002) of 0.26 L m"2 d"1, compares quite well to
the rate constant of 0.47 L m"2 d"1 reported by Butler
and Hayes (Butler and Hayes 2000).
4.3 Dioxane
1,4-Dioxane (1,4-Diethyleneoxide), often called
dioxane because the 1,2 and 1,3 isomers of dioxane
are rare, is a heterocyclic organic compound. It is a
colorless liquid with a faint sweet odor similar to that
of diethyl ether. It is classified as an ether and is used
as a solvent for fats, greases, and resins and in various
products including paints, lacquers, glues, cosmetics,
and fumigants. As a miscible compound, 1,4-Dioxane
is conservatively transported with no significant known
abiotic degradation pathway.
5. SUMMARY OF BIOTIC AND ABIOTIC
TRANSFORMATIONS
Chlorinated solvents are altered by intrinsic biotic
and abiotic processes. Transformations may be as such
that endpoints fall short of complete degradation to
innocuous compounds. The determination of which
endpoints are reached, the processes of transformation,
and the needed site data are critical for assessing and
modeling transport, and deciding on Monitored
Natural Attenuation (MNA) as a remedy. MNA is a
component of 22% of all Record of Decision (ROD)
in Superfund sites. Therefore, relevancy of MNA
research to OSWER and others in terms of reducing
uncertainty over field processes and better remedial
decision-making are the expected impacts of this work.
Many sites with chlorinated solvent contamination
may never proceed to a contaminant fate and transport
modeling stage, and therefore use the data to make
statistical inferences. For those sites, a thorough
recognition of transformation processes to form a
strong foundation for the development of a conceptual
site model and integrating site data to conceptualize
fate and transport processes without the benefit of
a computational model are essential. A quantitative
conceptual model, based upon transformation
knowledge and field observation provides the
framework for understanding and remediating a
site. The conceptual model also provides the basis
for developing and applying numerical models. This
document will briefly describe the process of applying
models (Section 6), given the uncertainty in processes
and input parameters. It will continue by discussing
alternative model formulations and their potential
utility. Transformation endpoints are summarized
below to facilitate in classification of observed plume
behaviors and patterns:
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 35
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Complete reductive dechlorination
PCE -» TCE -+ DCE
ethene, ethane
Incomplete/limited reductive dechlorination
PCE -> TCE -> DCE
or
PCE -» TCE -» DCE -* VC
PCE, 7~CE, DCE, VC plumes observed and all decreasing in
concentration, mass, and/or extent.
Ethane and ethene detected.
Strong reducing conditions (oxygen, nitrate, sulfate are depleted
relative to background wells)
Sufficient electron donor is present.
Requisite microbial community is present.
Presumed mechanism is reductive dechlorination.
PCE, TCE, DCE plumes or PCE, TCE, DCE, VC plumes
observed.
DCE and/or VC persist.
Weak reducing conditions (sulfate reduction and/or
methanogenesis is not occurring) and/or requisite microbial
community is not present.
Presumed mechanism is reductive dechlorination, but stopped
by lack of appropriate enzymes
Biotic/abiotic transformations
PCE -> TCE
PCE and TCE plumes observed and both decreasing in
concentration along the flow path.
No observed DCE or VC plumes observed.
Acetylene observed in ground water.
Strong reducing conditions (oxygen, nitrate, sulfate are depleted
relative to background wells)
Sufficient electron donor is present.
Requisite microbial community is present.
Mineralogical analysis would indicate presence of reactive
minerals
Presumed mechanisms are reductive dechlorination of PCE to
TCE, and abiotic transformation of PCE and TCE.
Degradation of TCA
1,1-DCE^CA
TCA, 1,1-DCA, and CA plumes are observed
Presumed mechanisms are reductive dechlorination of TCA to
DCA and abiotic transformation of PCE and TCE
No known culture has been found that is capable of complete
dechlorination of TCA to ethane.
Degradation of 1,4-dioxane
Presumed mechanism is aerobic respiration, as both growth-
supporting and non-growth supporting (i.e., cometabolism).
No evidence available to suggest abiotic transformation.
Monooxygenase reactions
:^M^^»»:
cytochrome p450
CO2
CO,
36 Ground Wah
Synthesis Report on State of Understanding of Chlorinated Solvent Transformation
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6. MODELING APPLICATIONS AND
CONCEPTUALIZATIONS FOR CHLORINATED
SOLVENT TRANSFORMATIONS
6.1 Historical Background
The scientific and conceptual basis for models of
ground-water flow and contaminant transport
date to the latter part of the 19™ century. Darcy's
experiment on flow through porous media had the
purpose of designing filters for the City of Dijon's
water supply (Darcy, 1856). It is important to note
that the experiment was performed on a sand filter,
where Darcy selected and prepared a relatively uniform
sand and placed it in an artificial environment: the
filter. Later the concept of how water flowed through
uniform materials was extrapolated to the natural
environment (Slichter, 1899), where an important
distinction holds: the materials are neither uniform
nor deliberately placed (for the most part). Methods
to quantify flow to wells (Thiem, 1932; Theis, 1935)
used mass conservation along with Darcy's Law
were developed in the early 20™ century. Although
simplified, these methods were successful for
determining flow to wells, largely because the location
of materials of differing conductivity is less important
for determining flow of water, than it is for transport
of contaminants, although this factor was not realized
at the time.
Mass conservation is also the main principle
underlying the transport of contaminants in aquifers.
Here the development of the transport theory in the
1950s (Bear, 1972) followed the development of
the theory of heat conduction in uniform materials
(Carslaw and Jaeger, 1959). In addition to the similar
basis in mass conservation, the original development
was for uniform materials. This limitation is
understandable because the numerical methods and,
more importantly, the computer power to solve
problems with heterogeneities did not exist at the time.
6.2 Types of Models
In the most commonly used approaches, the solutions
for ground-water flow and contaminant transport
are found separately. Thus the distinction is made
between ground-water flow models and contaminant
transport models. Although in this introduction both
are discussed, contaminant transport is the major focus
of this issue paper.
Two broad mathematical approaches have been
developed to solve the mass conservation equations
for ground-water flow and contaminant transport.
The first is the historic method of solving the partial
differential equation(s) for mass conservation. These
are exact solutions of the equations found through the
methods of calculus1. The solutions apply everywhere
throughout the domain, but require restrictive
assumptions. For contaminant transport, ground-
water flow must be steady (not varying with time) and
uniform (not varying with position). It is represented
as a simple constant in the analytic solution for
contaminant transport. Consequently heterogeneity
cannot be included, neither converging flow toward
wells nor irregular hydrologic boundaries such as
streams and rivers2.
The alternative is numerical solution which
approximates the solution over a set of points (usually
a grid), using approximate solution techniques for
the same partial differential equations. Numerical
methods are much more flexible than analytic
solutions because fewer major constraints are imposed.
This does not mean that the numerical methods are
not without limitations, but some of the basic and
severe constraints imposed on analytic solutions have
been overcome.
In the 1970s the first numerical models were
developed and made publically-available.
Concurrently there has been a parallel effort to develop
analytical models. Most developers justify the use of
these models as tools to test numerical models, a use to
which they are well suited, or as a screening tool. The
apparent idea behind screening tools is that because
they are simplified, they could be used to perform
quick analyses of transport when a full-blown analysis
is not warranted or possible. Caution is needed in the
1 Hybrid types have been developed that blur the distinction between
the two major types. Most familiar are the analytic element methods
which solve the ground-water flow equation analytically over a series
of domains, which are then linked to each other through what is
essentially a numerical approach.
2 Analytic element methods do have the ability to include irregular
boundaries, flow to wells and to a less common degree, heterogeneity.
Inclusion of these features would have been part of the motivation for
development of the method.
Synthesis Report on State of Understanding of Chlorinated Solvent Transformatio
Ground Water Issue 3 7
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use of simplified or screening models, and for example,
several questions need to be answered:
Are the assumptions in the simplified models met
by the field sites being screened?
Has it been demonstrated that simplified models
are appropriate for screening?
Have sufficient data been collected to support use
of the model? (i.e., to avoid a "garbage in/garbage
out" situation, have the sites been characterized)
Has the site-specific model (i.e., computer code plus
its site data) been shown to represent the specific
field site? If not, has an uncertainty analysis been
performed?
6.3 Parameter Measurement in the Field
Field methods exist to measure some model parameters
in the field, other parameters must be estimated. For
example, hydraulic conductivity can be measured by
aquifer pumping tests or slug tests. Aquifer pumping
tests might be impacted by rain or early termination
of pumping. Slug test results might be affected by
skin effects and the tests are acknowledged to provide
results close to the location of the tested well. In
contrast aquifer pumping test results can cover a wider
extent of the aquifer. Neither of these methods is
free from inaccuracies, nor do they typically produce
data that are as spatially refined as needed for detailed
simulation.
Other parameters are not directly measured. Porosity
is usually taken from literature values on aquifer type
and not determined on a site-specific basis. As will
be seen below, an approach to chlorinated solvent
modeling relies on first-order rate constants. These
are not directly measured but are estimated from
concentrations in wells across a site measured at
various times.
6.4 Model Application
Typical model applications use a combination of
measured, estimated and literature parameters as a
starting point. Even with the most comprehensive
investigation, numerical models could use more data
on parameter spatial variability than is available.
Because of limitations in the values for the initial
parameter values, parameters can be legitimately
changed to create a model that represents the field
data on contaminant concentrations. This process is
called calibration and is necessary to demonstrate that
the model reproduces conditions observed in the field.
Because it is essentially a process of interpolation, it
does not guarantee that the model will predict future
behavior, nor that the chosen parameters uniquely
determine the model results. Recent research on
calibration shows, in fact, that there is a limit beyond
which calibration cannot further refine parameters
towards reaching an ideal unique or "correct"
parameter set. This limitation derives from limitations
in the array of science that supports development and
application of models from the historic development
of the conceptual basis of the models, through field
measurement and application of computer codes.
6.4.1 Model Uses
What then are the best uses of models? There is
a near consensus that models are the best tool for
integrating the various processes occurring at field
sites. A consequence of applying the model can
be the understanding of which processes govern
transformation at a site. Questions can be asked
as: "Does abiotic transformation alone explains the
reduction in contaminant mass at this site?"
Recent writing on model application highlights the
limitations of presuming certainty from application
of environmental models in general. Oreskes (2003)
highlights the characteristics of problems where
application of environmental models is likely to be
highly successful. Two of her examples are planetary
motion, where predicted locations of planets can be
tested by nightly telescope observation, and weather
forecasting where the ability to forecast future weather
is known by all to be limited, but the forecasts are
valuable none-the-less. In a white paper published
by Ground Water, Konikow (2011) suggested that the
objectives for modeling be redefined.
Beyond understanding site behavior, models are
useful for situations where we plan to make future
measurements. Some examples are:
• More generally, design of remedial systems where
performance data will be collected to track the
progress of the remedy
• As a specific example: Prediction of the course of
38 Ground Water Issu
mthesis Report on State of Understanding of Chlorinated Solvent Transformation
-------
monitored natural attenuation (MNA) remedies,
where by definition monitoring will continue to
document the efficacy of remediation
6.4.2 Contaminant transport models
Fate and transport models are classified into two
categories:
A. Model with a sequential first-order decay
process
• Solute fate and transport model
• Sorption and retardation
• NAPL/water partitioning
• Groundwater flow velocity
• Biodegradation rate-constant
In the anaerobic reductive dehalogenation of
chloroethenes, chloroethenes were utilized as
respiratory electron acceptors. Bacteria can reductively
dechlorinate perchloroethene (PCE) to trichloroethene
(TCE), r»-dichlorothene (rw-DCE), vinyl chloride
(VC), and finally ethane (ETH). The ultimate
electron donor used in the process is H2 generated
from the fermentation substrates, often mediated by
mixed culture.
The sequential dechlorination is described in the
following pathway:
PCE -» TCE -» oj-DCE -» VC -» ETH
Each of the five solute (PCE and its daughter
products) simultaneous transport and degradation is
described by one-dimensional advection-dispersion
equation with first-order degradation kinetics. It
is assumed that the yield coefficients are based on
stoichiometric relations.
= L),
R
TCE
TCE
dt
dx
-v.
ac
TCE
dx
Y k r —
^TCE/PCE^PCE^PCE
r
^
dt
= D
dx2
-v
dC
DCE
V
1
DCE 'I TCE ^TCE
—k C
TCE ^DCE ^
VC
-V,
v
J
I- r
^ ^
VC/DCE ^DCE ^ DCE
dx
• v-T/r'^T/r
ETH
dt
dx
Y IT r —if r
-I. T7TTT If//"1 *vr/f ^—^ T/f *»/ ~C"~T~LJ ^—^
rLlrt/yL^ KC KC rLlrt
'ETH
concentrations (mg/L)
^PCE> ^TCE> ^DCE> ^vc> ^ETH " first-order degradation
rates (day"1)
YTCE/PCE> YDCE/TCE> YVC/DCE> YETHWC ~~ yield
coefficients (mg/mg)
Kinetic constants for the sequential degradation of
PCE
constant value (day"1)
kpcE
^TCE
I
-------
The model simulates the microbial transformation of
the seven solutes (PCE, TCE, DCE, VC, ETH, H2
and CH4), the growth and decay of three microbial
populations: PCE/TCE dechlorinators (dechl), DCE/
VC dechlorinators (dech2), and hydrogenotrophic
methanogens (meth) (or maybe homoacetogens). A
one-dimensional transport model is described with
dispersion, advection, and rate-limited sorption and
desorption, reductive dechlorination kinetics with
competitive inhibition and microbial growth and
decay.
oL,
TCE
a
*
C
res
f' 1 I
-^s,TCE,dechl L """
K
s,PCE,dechl
X
r -r
^ ^
, dechl
H2 ^ H2,th,dech)
L decftl
\
+ c
r,
3^ x>! 3x2 * 3x
/
^DCE,dech2^dech
C.
DCE
K
+c
DCE
x
C -C
^ ^
^s,H2,dech2
+ (€„ -C
_
~
9x 9jt
Cvc
V 1 | ^DCE
^S,VC,dech2\ L ' ^
V J^s,DCE,dech2 )
ch2-^ decK
\
/
r -
^
X
—C \
^H2,th,dech )
Beyond understanding site behavior, models are
useful for situations where we plan to make future
measurements. This Issue Paper will form the basis
for simulating chlorinated solvent transformation
along streamlines using biotic or abiotic processes as
appropriate.
40 Ground Water Issu
mthesis Report on State of Understanding of Chlorinated Solvent Transformation
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