EPA/600/A-96/123
Estimating the Changing Rate of Anaerobic Reductive Dechlorination of Chlorinated
Aliphatic Hydrocarbons in the Presence of Petroleum Hydrocarbons
David E. Moutoux, Leigh Alvarado Benson, Matthew A. Swanson, Todd H. Wiedemeier
Parsons Engineering Science, Inc.
John Lenhart
Division of Environmental Science & Engineering
Colorado School of Mines
John T. Wii son
National Risk Management Research Laboratory
US Environmental Protection Agency
Jerry E. Hansen
Air Force Center for Environmental Excellence
Abstract
Recent laboratory and field results demonstrate that different chlorinated aliphatic
hydrocarbons (CAHs) ultimately can be transformed into innocuous chemical compounds in
many aquifer systems. Transformation can be the result of either cometabolic reactions or
reduction oxidation (redox) reactions. In the latter, chlorinated compounds such as
tetraehloroethene (PCE), trichloroethene (TCE), and dichloroethene (DCE) can be used in
anaerobic, reducing aquifers as alternate electron acceptors during the oxidation of organic
matter such as the petroleum hydrocarbons benzene, toluene, ethylbenzene, and xylenes (BTEX).
When used as electron acceptors, these CAH compounds are sequentially dechlorinated into less
chlorinated compounds such as vinyl chloride (VC) or ethene. While the rate of petroleum
hydrocarbon degradation is generally reasonably approximated as a first-order process, it is not
clear that the degradation of electron acceptors such as CAH compounds are best estimated in a
similar manner. Because the rate of reductive dechlorination of CAHs in the presence of
petroleum hydrocarbons is a function of both the concentration of the petroleum hydrocarbons
and the concentration of the electron acceptors themselves, a second-order kinetic model may
provide a better approximation of the rate of BTEX/CAH degradation over time and distance.
Several methodologies may be used to estimate the rate of reductive dechlorination of CAH
compounds when they are being used to oxidize BTEX compounds. Both first-order and second-
order approximations of CAH degradation rates can be useful in predicting CAH/BTEX plume
behavior, although the second-order approximation will be especially useful in determining
whether the system will first exhaust the available supply of electron donors (i.e., starve) or
electron acceptors (i.e., strangle).
Introduction
Numerous laboratory and field studies have shown that microorganisms indigenous to the
subsurface environment can degrade a variety of fuel hydrocarbons and CAHs (e.g., Chapelle,
1993; Young and Cerniglia, 1995). During biodegradation, microorganisms transform available
nutrients into forms useful for energy and cell reproduction by facilitating the transfer of
electrons from donors to acceptors. This results in oxidation of the electron donor and reduction
of the electron acceptor. Electron donors include natural organic material and fuel hydrocarbons
such as the BTEX compounds. When fuel hydrocarbons are utilized as the primary electron
donor for microbial metabolism, these compounds are typically completely degraded or
detoxified (Bouwer, 1992). Electron acceptors are elements or compounds that occur in
relatively oxidized states. Important inorganic electron acceptors commonly found in
groundwater include dissolved oxygen (DO), nitrate, manganese, ferric iron, sulfate, and carbon
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dioxide. The reduction of these inorganic electron acceptors during fuel hydrocarbon oxidation
causes measurable changes in groundwater chemistry. The reduction in concentration of
inorganic electron acceptors along groundwater flowpaths can be used qualitatively to document
biodegradation (Wiedemeier et al., 1995) and quantitatively to estimate degradation rates
(McAllister and Chiang, 1994; Wilson et al, 1994; Chapelle et al, 1996).
In addition to these common inorganic electron acceptors, recent laboratory and field data
suggest that CAH compounds may serve as alternate (organic) electron acceptors when sufficient
organic substrate is available in the system (e.g., Gossett and Zinder, 1996). The presence of
elevated concentrations of organic substrate in the aquifer (e.g., from a fuel spill) results in a
surplus of electron donors, effectively increasing the reducing potential of the groundwater. If an
appropriate succession of biological mediators is available to facilitate redox reactions between
surplus electron donors and electron acceptors, the more common inorganic electron acceptors
such as DO, nitrate, and ferric iron will be reduced (depleted). Once the oxidizing potential of
the system has been sufficiently reduced to allow sulfate reduction, the microorganisms may
facilitate equally thermodynamically favorable redox reactions between the remaining surplus
electron donors and alternate (organic) electron acceptors such as the CAH compounds. During
these reactions, the fuel hydrocarbons (electron donors) will be completely oxidized to water and
carbon dioxide, and the CAH compounds (electron acceptors) will be sequentially dehalogenated
to less chlorinated compounds such as ethene.
Consequently, the effectiveness of natural chemical attenuation processes at minimizing
downgradient migration and eventually eliminating contaminant mass depends on the rates of
microbial degradation processes (National Research Council, 1993; Salanitro, 1993; Wiedemeier
et al, 1995, Chapelle et al, 1996). The use of substrates such as organic electron donors and
electron acceptors by microorganisms for microbial growth generally follows enzyme saturation
(Monod and variations on Monod) kinetics, which is derived from the Michaelis-Menten enzyme
kinetics equation:
v =
VmaxSA'
VKs + S
0)
where v is the rate of substrate uptake (moles/time), vmax is the maximum rate of substrate uptake
(moles/time-gram cells), Kj is the concentration of substrate at which v = 0.5 vmax (moles/L), S is
the concentration of substrate (moles/L), and X is the mass of cells (grams). When the uptake of
a substrate is limited by enzyme availability (e.g., at low substrate concentrations, S«Ks), and
when the microbial mass is neither increasing or decreasing with time (i.e., growth = loss), the
Michaelis-Menten equation can be approximated by first-order kinetics:
v ~ kS (2)
where k is a rate constant (time'1). The first-order reaction rate dependence has implications on
degradation rates at low concentrations: the reaction rates become progressively slower with
decreasing substrate concentration.
First-order kinetics cure believed to reasonably approximate microbial degradation of BTEX
in aquifers because the system is generally assumed to be limited by one substrate, the organic
electron donor. This generally is a reasonable assumption when only inorganic electron
acceptors, which can be infinitely supplied from upgradient groundwater or infiltrating
precipitation migrating into surplus electron donor mass, are being used in coupled redox
reactions. The first-order assumption also simplifies fate and transport modeling of a complex
phenomenon. Once the effects of nondestructive attenuation processes such as advection,
dispersion, and sorption on microbially reactive compounds in groundwater are quantified, the
apparent mass loss of BTEX along groundwater flowpaths at several sites have been shown to be
reasonably approximated by a first-order degradation rate constant (McAllister and Chiang,
1994; Buscheck and Alcantar, 1995; Wiedemeier et al, 1995 and 1996a; Chapelle et al, 1996).
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Use of the first-order approximation also may be appropriate to estimate rates of
dechlorination of CAH compounds if it is assumed that the utilization of the primary substrate
(electron donor or BTEX) is not dependent on the presence of CAH compounds (electron
acceptor). This may be a reasonable assumption if sufficient electron donor mass is available
(i.e., significant BTEX concentrations) or if the microbial population capable of facilitating
redox reactions between BTEX and CAH is small and exhibiting low growth rates relative to
other microbial populations. However, the use of first-order kinetics may not be appropriate
when more than one substrate is limiting microbial degradation rates or when microbial mass is
increasing or decreasing (i.e., M). Because the oxidation of BTEX depends on the reduction of
inorganic electron acceptors or CAH compounds, these types of redox reactions can be defined
as bimolecular reactions
where k is a second-order rate constant (mass/time), and the disappearance (or degradation) of B
is first-order in both [B] and [A], the electron acceptors). This means that the rate of
degradation of B (BTEX) depends both on the concentration of B (BTEX) and the concentration
of A (electron acceptors). Conversely, the same is true of [A]: the degradation of the electron
acceptors is first-order in both [B] and [A], Intuitively, this makes sense in terms of how these
types of degradation reactions occur: if insufficient electron acceptor mass is available to oxidize
the fuel hydrocarbon mass, the reaction will stall (i.e., the system will strangle); if insufficient
electron donor mass is available, degradation will cease (i.e., the system will starve). The former
is important in terms of predicting the maximum downgradient migration and persistence of a
dissolved BTEX plume. The latter may be extremely important for mixed BTEX/CAH plumes
to determine whether "natural" reductive dechlorination processes can minimize CAH migration
and achieve complete dechlorination before complete oxidation (degradation) of BTEX occurs.
The remaining sections of this paper presents several methodologies for estimating both
first-order CAH dechlorination rates and second-order BTEX/CAH degradation rates. Site data
from several groundwater plumes, collected as part of a nationwide project sponsored by the Air
Force Center for Environmental Excellence, are used to illustrate these different methods. A
comparative analysis of the predicted degradation rates and their implications on the remedial
decision-making process also is presented.
First-Order Decay Estimation Methods
The change in a solute's concentration in groundwater over time often can be estimated
using the linear form of the first-order kinetic model:
where C = concentration at time t [Hg/L], C0 = source concentration [|ig/L], k = overall
attenuation rate (first-order rate constant) [day" ]
Using this relationship, a total first-order attenuation rate can be estimated for a site by
producing a log-linear plot of total contaminant concentration versus travel time for a series of
points along the flowpath of a contaminant plume. If the data plot along a straight line, the
relationship is first-order, and an exponential regression analysis can be performed, with the
slope of the resulting equation equivalent to the total attenuation rate. This overall attenuation
rate groups all processes acting to reduce contaminant concentrations and includes advection,
dispersion, dilution from recharge, sorption, volatilization, and biodegradation. This approach is
A + B —> product(s)
where the rate of change of [B], assumed to be the electron donor, is
k[B][A]
(3)
C = C0e"*'
(4)
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very sensitive to hydrogeologic variability, as well as to the proximity of the sampled points to
the dominant flow path of the plume, both laterally and vertically. This can contribute to the
generation of less than desirable correlation coefficients (R ) and bring the first-order assumption
into question.
Total chlorinated ethene attenuation rates have been estimated for three sites: a former fire
training area at Plattsburgh Air Force Base (AFB) (FT-002), a former fire training area at Cape
Canaveral Air Station (AS) (FT-17), and a former bomber assembly plant at Offutt AFB (Bldg.
301). A brief summary of historical site information is provided in Table 1; site characterization
data are summarized in Table 2. Total estimated chlorinated ethene attenuation rates for the
three sites (Table 3) ranged from 0.00021 to 0.00051 day"1, with the magnitude of the rates
closely tied to the average retarded contaminant velocity. As an example, a log-linear plot of
data collected from Plattsburgh AFB in 1995 is provided in Figure 1.
Estimating First-Order Biodegradation for a Steady-State Plume
In order to ensure that some portion of observed decreases in contaminant concentrations
can be attributed to biodegradation, measured contaminant concentrations must be corrected for
the effects of dispersion, dilution, and sorption. Buscheck and Alcantar (1995) derive a
relationship that allows calculation of first-order biodegradation rate constants for steady-state
plumes. This method involves coupling the regression of contaminant concentration (plotted on
a logarithmic scale) versus distance downgradient (plotted on a linear scale) to an analytical
solution for one-dimensional, steady-state, contaminant transport that includes advection,
dispersion, sorption, and biodegradation.
The relationship developed by Buscheck and Alcantar (1995) was applied to the data from
all three sites, although each of these plumes is suspected to be expanding. As expected, the
rates attributed to biodegradation are less than the total attenuation rates, with estimated rates
ranging from one-half to three-quarters of the total attenuation rate. Data and results are
presented in Tables 2 and 3, respectively. Figure 2 provides, as an example, the log-linear plot
used in the calculation of the 1995 Plattsburgh AFB biodegradation rate. Because this technique
uses the same concentration data as the total attenuation technique, it is equally sensitive to
sampling locations and hydrogeologic variability.
For an expanding plume, this first-order approximation can be viewed as an upper bound on
the biodegradation rate. Use of this method results in an overestimation of the rate of
biodegradation because a typical expanding plume exhibits decreasing source area
concentrations, increasing downgradient (and crossgradient) concentrations, or both. Over time,
these changes result in a decreasing slope on a log-linear plot, and consequently a decreasing
biodegradation rate.
Estimating First-Order Reductive Dechlorination: the Carbon Core as a Tracer
A convenient way to isolate the rate of biodegradation from other attenuation processes is to
use as tracers compounds or elements associated with the contaminant plume that are relatively
unaffected or predictably affected by biological processes occurring within the aquifer. When
present, the trimethylbenzene isomers associated with fuels can serve as useful tracers under
certain geochemical conditions (Wiedemeier et ai, 1995 and 1996a). Likewise, chloride, a
degradation product of chlorinated solvent biodegradation has the potential to serve as a useful
tracer (Wiedemeier et al., 1996b). This section describes a tracer method that can be used with
reductively dehalogenated solvent plumes, and involves tracking the "carbon" core of the
chlorinated compounds in relation to the remaining chlorine mass.
Measured tracer and contaminant concentrations from a minimum of two points along a flow
path can be used to estimate the amount of contaminant remaining at each point if biodegradation
had been the only attenuation process operating to reduce contaminant concentrations. To
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accomplish this, it is assumed that the fraction of contaminant remaining as a result of all
attenuation processes is equivalent to the fraction of contaminant remaining as a result of non-
destructive attenuation mechanisms only, multiplied by the fraction of contaminant remaining as
a result of biodegradation. The fraction of contaminant remaining as a result of all attenuation
processes can be computed from the measured contaminant concentrations at two points along a
flow path. The fraction of contaminant remaining as a result of non-destructive attenuation
mechanisms only can be estimated from the tracer concentrations at the same two points, because
an ideal tracer is affected by non-destructive attenuation mechanisms to the same degree as the
contaminant of interest and is not affected by biologic processes. The following equation uses
these assumptions to solve for the estimated downgradient contaminant concentration if
biodegradation had been the only attenuation process operating between two points (/ and i-l)
along the flow path:
C - C
^ /VIW
t,corr
{ Ci\
{1
\
-1
1 ,corr
n
r»
\J
\ 1
/ J
(5)
where Cicpr = corrected contaminant concentration at point i; C,./ corr= corrected contaminant
concentration at point i-l (Note that if point i-l is the first or most upgradient point, C;.IjCorT is
equivalent to the observed contaminant concentration.); C, = observed contaminant concentration
at point i; C,. ,= observed contaminant concentration at point i-l; T,- = observed tracer
concentration at point i; and T,_ 7= observed tracer concentration at point i-l.
This equation can be used to estimate the theoretical contaminant concentration resulting
from biodegradation alone for every point along a flow path on the basis of the measured
contaminant concentration at the point of plume origin and the contaminant/tracer ratios between
consecutive points along the flow path. This series of points can then be used to estimate a first-
order rate of biodegradation as described for estimating total attenuation rates.
During reductive dechlorination, the source chlorinated solvent undergoes successive
transformations involving the replacement of a chlorine atom by a hydrogen atom; however, the
carbon core of both the parent and daughter compounds remains unchanged (i.e., no carbon
bonds are broken). The carbon core is subject to the same non-destructive attenuation
mechanisms that act on the larger chlorinated molecule, but it is unaffected by biologically
mediated reductive dechlorination. For this reason, tracking the carbon core of dissolved
chlorinated solvents can serve as a theoretically perfect "tracer" for biodegradation via reductive
dechlorination.
In order to use the carbon core of the chlorinated parent and daughter compounds as a
"tracer" for reductive dechlorination, "equivalents" for the dissolved mass of carbon and chlorine
must be calculated for each point along a flow path. The "equivalents" are calculated by first
converting contaminant concentrations into molar concentrations. For chlorinated ethenes, the
carbon equivalent is calculated by multiplying the number of carbon atoms per molecule of
chlorinated ethene (2) by the sum of the molar concentrations for PCE, TCE, DCE, VC, and
ethene:
Ce'qj = 2 (MpcEj + •^TCE.i + M)CE,i + ^VC,i + Athene,i) (6)
where Ceq; = carbon equivalent at point i; A/PCE i = molar concentration of PCE at point i; MTCE i
= molar concentration of TCE at point i; A/dce,i = molar concentration of DCE at point i; MVC i =
molar concentration of VC at point i; and A/nthene.i = molar concentration of ethene at point i.
The chlorine "equivalent" is defined as the sum of the products of molar concentration and
chlorine atoms per molecule for each parent and daughter compound. For the chlorinated
ethenes, the numbers of chlorine atoms per molecule are 4 for PCE, 3 for TCE, 2 for DCE, 1 for
VC, and 0 for ethene:
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Cleqj — (Mpce ;*4) +• (MTCE>i*3) + ;*2) + Mvc.i
(7)
where Cleqj = chlorine equivalent at point i.
Using equation 5, and substituting Ceq for tracer concentrations and Cleq for observed
contaminant concentrations, yields the theoretical total CAH concentrations at downgradient
locations if reductive dechlorination had been the only natural attenuation process operating
along the flow path. The same process can be used to determine the theoretical chlorine
equivalents. Chlorine equivalents, carbon equivalents, the corrected total CAH concentrations,
and the corrected chlorine equivalents for the Cape Canaveral AS, Plattsburgh AFB, and Offiitt
AFB sites are presented in Table 4. Thexorrected CAH concentrations are useful for comparison
to other techniques; the corrected chlorine equivalents simplify visualization of the reductive
dechlorination rate. Either the corrected total CAH concentrations or corrected chlorine
equivalents can be used to calculate identical first-order rates for dechlorination (Table 3). An
example log-linear plot is provided in Figure 3 for the 1995 Plattsburgh AFB calculation.
The results serve to illustrate two important aspects of this technique. First, the calculated
first-order rate is for reductive dechlorination only. The Bldg. 301 plume at Offiitt AFB is
characterized by predominantly aerobic conditions and low daughter product concentrations
throughout large portions of the plume; therefore, reductive dechlorination is expected only in
isolated portions of the plume. This technique estimates a low reductive dechlorination rate with
a low R because limited reductive dechlorination appears to be occurring both at the head and
the tail of the plume; however, little to no reductive dechlorination occurs through the central
portion of the plume. Anaerobic, reducing conditions with large daughter product concentrations
prevail at the fire training areas at Cape Canaveral AS and Plattsburgh AFB. Consequently, both
have reductive dechlorination rates estimated with a high degree of correlation.
Secondly, the rate estimate does not adequately assess the total biodegradation rate if
biodegradation mechanisms other than reductive dechlorination are operant. Alternate
biodegradation avenues are available for lower molecular weight solvents such as VC as
groundwater conditions become less reducing. For instance, at the Plattsburgh AFB FT-002 site,
groundwater geochemistry becomes less reducing between 2,000 and 2,500 feet downgradient
from the source area; therefore, a reductive dechlorination rate cannot be calculated beyond this
point. The combination of slowing reductive dechlorination rates and the destruction of VC (and
perhaps other parent and daughter products) by alternate biodegradation processes renders the
technique inappropriate.
Second-Order Degradation Rate Estimates
Although a first-order rate assumption may provide a reasonable approximation of how
BTEX and CAH compounds are degrading in groundwater systems, this approach may neglect
the importance of the electron donor-electron acceptor redox couples or the variable rate of
biomass growth expected throughout the plume. As discussed previously, a first-order kinetic
model may not provide the best approximation of how CAH compounds are dechlorinated
(biodegraded) in the presence of another limited substrate, the electron donor (BTEX). Because
highly-chlorinated CAH compounds are rarely used as primary substrates for microbial
metabolism (e.g., McCarty and Semprini, 1994), the dechlorination of these compounds is
dependent upon the microbial utilization of a primary substrate such as BTEX. Therefore, the
degradation kinetics of this dual-dependency reaction may be more appropriately approximated
by a bimolecular reaction rate expression (see equation 3). The linear form of this second-order
equation is:
In
[A]o + [B]o
[A]o[B]
.[B]o[Al
= kt
(8)
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where [B]0 = initial concentration of electron donor (jxg/L); [B] = measured concentration of
electron donor (fig/L); [A]c = initial stoichiometry-normalized concentration of electron acceptor
(p.g/L); [A] = measured stoichiometry-normalized concentration of electron acceptors (ng/L),
and k = second-order degradation rate constant (L/p.g-day). Table 5 presents the stoichiometry-
normalized concentrations of these electron acceptors measured at the Plattsburgh AFB FT-002
site, which were used to estimate second-order rate constants. The reason that stoichiometry-
normalized electron acceptor concentrations were used to develop second-order reaction rate
constants was to develop a weighted estimate of oxidizing potential.
If the objective of the second-order approximation is to estimate the rate of CAH
degradation in the presence of petroleum hydrocarbons, no significant concentration correction - -
for the effects of nondestructive attenuation processes (i.e., advective-dispersive transport and
sorption) is necessary. These nondestructive attenuation processes will have the same general
concentration-reducing effects on anthropogenic organic electron acceptors (CAH) and
anthropogenic organic electron donors (BTEX). Both of these types of compounds will be
advectively transported and dispersed from the source area, and are subject to sorption.
Although the sorptive characteristics of the BTEX compounds are slightly greater than that of the
CAH compounds, the organic carbon content of the saturated soils at some sites (e.g.,
Plattsburgh) may be low enough to minimize significant retardation. In contrast, if the objective
of the second-order approximation was to estimate the rate of degradation of anthropogenic
electron donors (BTEX) in the presence of only natural electron acceptors (i.e., the common
inorganic electron acceptors), the concentration-reducing effects of nondestructive attenuation
processes on the BTEX compounds would have to be considered. To simplify this example, this
paper investigates only the second-order relationship between CAHs and BTEX at Plattsburgh,
where limited contaminant sorption is expected.
Figures 4 and 5 present the second-order reaction rate constants derived for BTEX-CAH
redox reactions for the Plattsburgh AFB FT-002 site (1993 and 1995 sampling events). The
second-order reaction rate constant is equal to the slope of the best fit line as a function of the
second-order linear expression and contaminant transport time (days). Although the coefficients
of correlation (R) indicate that the data are reasonably approximated by the second-order
relationship, the correlation is not as strong as suggested by the first-order methods. The lack of
absolute correlation may be a function of limiting the reaction dependency to only anthropogenic
chemicals, rather than accounting for the effects of inorganic electron acceptors known to be
involved in redox reactions at the site. Additionally, the lower correlation may be a reflection of
the slight difference between the effects of nondestructive attenuation processes on the electron
donors and electron acceptors.
These estimated second-order rate constants can then be used to estimate the changing
degradation rates for redox reactions involving CAH compounds in the presence of BTEX as a
function of distance from the source and/or time as follows:
degradation rate (fig/L-day)= k[B][A] (9)
Table 5 also presents the calculated degradation rates at different points along the
groundwater flowpath and for different sampling events. As these estimates indicate, the rate of
degradation is fairly high when sufficient concentrations of both electron donors and electron
acceptors are present. This result is consistent with the measured chloride concentrations, that
suggest dechlorination is occurring at and immediately downgradient from the core of the
BTEX/CAH plume (Table 5). However, as the concentrations of each diminish, the rate of
degradation decreases. Notably, the estimated degradation rate at the Plattsburgh FT-002 site
during the 1995 sampling event was lower than that estimated from the 1993 sampling data.
These results suggest the rate of degradation is changing both as a function of distance along the
plume flowpath and as a function of time.
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These results may be significant in terms of anticipating whether CAH plumes can be
"naturally" dechlorinated to less chlorinated compounds such as ethene. When the concentration
of the electron donor has been sufficiently oxidized and/or when the electron acceptor "out
migrates" the electron donor, the rates of degradation will decrease and possibly even cease. For
example, at the Pittsburgh FT-002 site this means that as the CAH mass migrates beyond the
available electron donor mass (see Table 5), the rate of degradation becomes significantly less
than that predicted to be occurring in the core of the plume. Unfortunately, these results mean
that the remaining CAH mass may not be subject to further degradation via reductive
dechlorination processes. Degradation may still occur, however, if the remaining CAH mass is
amenable to oxidation reactions (e.g., vinyl chloride).
Discussion
A comparative analysis of the results of these methodologies is warranted. For the
Pittsburgh AFB FT-002 site, the biodegradation rates estimated using a first-order assumption
are similar to the average rate estimated using the second-order approximation (see Tables 3 and
5). The average CAH reductive dechlorination rate estimated using the first-order approximation
based on the 1993 sampling event is 0.00011 day" (which corresponds to a half-life of about 17
years). The average BTEX-CAH biodegradation rate estimated using the second-order approach
is about 0.00014 day"1 (which corresponds to a half-life of 13.5 years). Although these methods
yield apparently comparable average biodegradation rates, the second-order approximation
clearly shows how the rate of biodegradation of BTEX-CAH can vary within a plume, both as a
function of location and time. Specifically, the second-order biodegradation rate estimated for
the plume core at the Pittsburgh AFB FT-002 site from the 1993 sampling data is almost twice
the estimated average biodegradation rate. This results in a half-life of about 8 years within the
plume core. As the electron donors and electron acceptors migrate away from the source area,
the biodegradation rate decreases due to the rate-limiting availability of both compounds. For
example, by the time the measured BTEX decreases by two orders of magnitude from the source
area concentration (i.e., at well MW-02-042), the biodegradation rate has decreased by more than
two orders of magnitude. This effectively increases the half-life of remaining mass in the part of
the plume to about 900 years! Of course, other destructive attenuation mechanisms such as
oxidation processes (rather than reductive processes) may play a large role in attenuation at this
point.
Similar relationships are illustrated by comparing the 1993 data sets and the 1995 data sets.
The first-order biodegradation rate based on 1995 sampling results is about half the
biodegradation rate estimated from 1993 sampling results. The difference in the second-order
biodegradation rates between these two sampling events is even more marked. In the source
area, the rate of biodegradation estimated in 1995 was only about 37 percent of the rate estimated
in 1993. This reduction in biodegradation rate may be attributable to the availability of the
electron donor. In 1993, the proportion of electron donor mass to organic electron acceptors
decreased by about 38 percent between 1993 and 1995. This suggests that the reduction in
electron donor mass may become rate-limiting over time, even in the source area.
These biodegradation rate estimate methodologies provide valuable information that should
be factored into the remedial decision-making process. The first-order biodegradation estimates
provide a good indicator of observed dechlorination at a single point in time as averaged across a
plume flowpath. However, the second-order degradation estimates show that biodegradation
rates are variable, which can have a profound effect on the anticipated effectiveness of natural
chemical attenuation processes and the need to implement additional engineered actions or
exposure controls.
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Proceedings, Symposium on Intrinsic Bioremediation of Ground Water, August 30-
September 1,1994, Denver, CO, pp. 60-72.
Young, L.Y. and Cerniglia, C.E., eds., 1995, Microbiological Transformation and Degradation of
Toxic Organic Chemicals: Wiley-Liss, New York.
Contributing Author Information
Brief biographical sketches and mailing addresses for contributing authors are provided below.
Mr. David E. Moutoux is an environmental engineer and project manager with Parsons Engineering Science, Inc.
[1700 Broadway, Suite 900, Denver, CO 80290, (303) 831-8100], His responsibilities include site characterization,
geochemical analysis, and contaminant transport modeling to investigate/document natural attenuation of dissolved
petroleum and chlorinated aliphatic hydrocarbons. He holds an A.B. in Earth Science and Engineering, a B.E. in
Engineering Science, and an M.E. in Engineering Science, all from Dartmouth College.
Ms. Leigh Alvarado Benson is an environmental engineer/risk assessor with Parsons Engineering Science, Inc.
[1700 Broadway, Suite 900, Denver, CO 80290, (303) 831-8100], Her responsibilities include the development of
risk-based corrective actions, quantitative exposure assessments, probabilistic risk assessments, and geochemical
analysis. She holds a B.S. in Zoology (chemistry emphasis; anthropology second major) from Duke University, and
an M.S. in Environmental Engineering from the Colorado School of Mines.
-------
Mr. John Lenhart is a doctoral candidate in the Division of Environmental Science & Engineering at the Colorado
School of Mines [Golden, CO 80401, (303) 384-2116]. His current research is focused on the role of natural
organic material in transporting trace chemicals in natural water systems. He holds a B.S. in Mechanical
Engineering, and an M.S. in Environmental Engineering from the Colorado School of Mines.
Mr. Matthew A. Swanson is a hydrogeologist with Parsons Engineering Science, Inc. [1700 Broadway, Suite 900,
Denver, CO 80290, (303) 831-8100]. His responsibilities include site characterization, geochemical analysis, and
groundwater flow and contaminant transport modeling to investigate/document natural attenuation of petroleum and
chlorinated aliphatic hydrocarbons. He holds a B.S. in Geology from Trinity University, and an M.S. in
Hydrogeology from the University of Wisconsin-Madison.
Mr. Todd H. Wiedemeier is a principal hydrogeologist and technical director with Parsons Engineering Science, Inc.
[1700 Broadway, Suite 900, Denver, CO 80290, (303) 831-8100]. He has been working with several research
organizations, the Air Force, and the Navy to develop and field test sampling procedures, analytical methodologies,
and data evaluation techniques to scientifically document field-scale in situ biodegradation of petroleum and
chlorinated aliphatic hydrocarbons. He is primary author of several technical "how-to" protocols on how to
investigate the effects of natural attenuation processes on dissolved organic contaminants. He holds a B.S. in
Geology from Colorado State University, and an M.S. in Geology from Wichita State University.
Dr. John T. Wilson is a senior research microbiologist with the US Environmental Protection Agency's National
Risk Management Research Laboratory (NRMRL) [919 Research Drive, Ada, OK, (405) 332-8800]. Dr. Wilson's
current research interests involve the study of microbial processes that determine the fete and transport of organic
contaminants in the subsurface environment. He has more than 12 years of extensive experience in site
characterization, process definition, and modeling of chemical attenuation processes. He received a Ph.D. in
Microbiology from Cornell University.
Mr, Jerry E. Hansen is a program manager with the Air Force Center for Environmental Excellence at Brooks Air
Force Base [2504 D Drive, Suite 3, Brooks AFB, TX 78235, (210) 536-4353], Mr. Hansen is currently managing
AFCEE's innovative remedial approach/technology efforts at more than 40 Air Force facilities nationwide. He
holds a B.S. in Mechanical Engineering from the University of Nebraska, and an M.S. in Aeronautical Engineering
from the University of Texas-Austin.
-------
TABLE 1
SUMMARY OF SITE INFORMATION
Sampling
Location
Historical Use
Contamination
Type and
Concentration
Age of
Release
Length of Plume
Hydrogeologic
Parameters
Cape Canaveral
Fire training and drum
storage area operating
between 1965 and 1985.
Comingied
waste fuels and
solvents
Chronic release
from 1965 to
1985.
Approximately 1,200
feet to surface water
discharge point
Groundwater Velocity: 0.27 ft/day
Contaminant Velocity: 0.16 ft/day
Dispersivity: 130 feet
Plattsburgh
Fire training area
operating between
1950's and 1989.
Comingied
waste fuels and
solvents
Chronic release
from the 1950's
to 1989.
More than 3,100 feet
Groundwater Velocity: 0.39 ft/day
Contaminant Velocity: 0.30 ft/day
Dispersivity: 300 feet
Offutt
Bomber assembly from
1944 to 1945; guided
missile assembly from
1959 to 1965.
Solvents used
in degreasing
operations
Unknown
Approximately 3,000
feet
Groundwater Velocity: 0.3S ft/day
Contaminant Velocity: 0.29 ft/day
Dispersivity: 250 feet
TABLE 2
SUMMARY OF SITE CHARACTERIZATION DATA
Distance
Predicted CAH
Total
Ethene &
Total PCE, TCE,
Sampling
from Source
Travel Time
PCE
TCH
1,2-DCE
VC
Ethane
DCE, & VC
Location
(feet)
(days)
(Hg/L)
(Hg/L)
(Hg/L)
(Hg/L)
(Pg/L)
(Hg/L)
Cape Canaveral
CCFTA2-9S
0
0
56
15,800
98,889
3,080
30
117,825
MP-3
775
4,880
BDL
220
3,502
3,080
243
6,802
CPT-4
885
5,572
BDL
16.5
781.9
797
BDL
1,595
MP-6
1,145
7,209
BDL
24.3
1,216
2,520
120
3,760
CCFTA2-14
1,270
7,996
BDL
42
996
6,520
228
7,558
MP-4S
1,320
8,311
BDL
19
575
5,024
153
5,618
Plattsburgh
MW-02-014 (1993)
0
0
BDL
1,030
9,050
4.5
3.2
10,085
84-M (1993)
460
1,523
BDL
BDL
1,320
1,050
28.4
2,370
MW-02-019 (1993)
730
2,417
BDL
1.9
3,540
384
39
3,926
MW-02-042 (1993)
2,600
8,607
BDL
98.5
1,570
2.2
3.5
1,671
MW-02-043 (1993)
3,750
12,414
BDL
373
10.6
BDL
BDL
384
Point A (1995)
0
0
BDL
25,280
51,412
BDL
BDL
76,692
Point B (1995)
970
3,211
BDL
2
14,968
897
35
15,867
Point C (1995)
1,240
4,105
BDL
3
10,035
1,430
182
11,468
Point E (1995)
2,560
8,474
BDL
24
2,218
8
BDL
2,250
Point F (1995)
3,103
10,272
BDL
1
226
5
BDL
232
Point A (1996)
0
0
BDL
580
12,626
BDL
BDL
13,206
Point B (1996)
970
3,211
BDL
1
9,376
1,520
13
10,897
Point C (1996)
1,240
4,105
BDL
I
10.326
1,050
170
11,377
Point D (1996)
2,050
6,786
BDL
BDL
1,423
524
4
1,947
Point E (1996)
2,560
8,474
BDL
17
1,051
12
BDL
1,080
Point F (1996)
3,103
10,272
BDL
BDL
177
4
BDL
181
Offutt
MW7I
0
0
BDL
17,500
1,235
BDL
BDL
18,735
MW18
420
1,461
BDL
3,610
4
BDL
BDL
3,614
MW14
510
1,774
BDL
2,940
11.5
BDL
BDL
2,952
MW8I
980
3,409
BDL
201
BDL
BDL
BDL
201
MW9S
1,460
5,078
BDL
372
23
1
BDL
396
TWI
2,020
7,026
BDL
438
58
BDL
BDL
496
MWI1
2,520
8,765
BDL
15.2
77.3
BDL
BDL
92.5
-------
TABLE 3
ESTIMATED FIRST-ORDER ATTENUATION RATES
RJ for Total
Steady-State
Attenuation
Total
Plume
and Steady-
Reductive
R* for
Attenuation
Biodegradation
State Plume
Dehalogenaiion
Reductive
Sampling Location
(day"1)
(day1)
Biodegradation
(day4)
Dehalogenation
Cape Canaveral
January 1996
0.00036
0.00026
0.59
0.000085
0.97
Ptattsburgh
December 1993
0.00021
0.00017
0.84
0.000112
1.0
August 1995
0.00052
0.00025
0.95
0.000063
0.98
May 1996
0.00042
0.00025
0.86
0.000033
0.99
Qffutt
June 1996
0.000S1
0.00029
0.75
0.000029
0.53
TABLE 4
FIRST-ORDER REDUCTIVE DEHALOGENATION RATE CALCULATION
Carbon
Chlorine
Corrected
Carbon
Chlorine
Corrected j
Sampling
Equivalents
Equivalents
CAH
Sampling
Equivalents
Equivalents
CAH
Location
L)
(HM/L)
(Hg/L)
Location
(VM!L)
(Hg/L) I
Cape Canaveral
Ptattsburgh
MW-02-014 (199
1,446
1,638
10,088
CCFTA2-9S
2,382
2,452
117,799
84-M (1993)
340
323
8,462
MP-3
191
127
75,647
MW-02-019 (199
266
230
7,703
CPT-4
42
29
79,956
MW-02-042 (199
46
46
8,913
MP-6
115
66
65,849
MW-02-043 (199
5
5
8,769
CCFTA2-14
246
126
58,526
MP-4S
184
93
57,704
Point A (1995)
1,446
1,638
76,692
Point 8(1995)
340
323
64,332
Offutt
Point €(1995)
266
230
58,560
Point E (1995)
46
46
67,762
MW71
292
425
18,735
Point F (1995)
5
5
66,670
MW18
55
83
19,287
MW14
45
67
19,263
Point A (1996)
269
274
13,206
MW8I
3
5
19,296
Point 8 (1996)
243
218
11,644
MW9S
6
9
18,735
Point C (1996)
259
230
11,541
TW1
8
11
18,318
Point D (1996)
46
38
10,566
MW1I
2
2
13,679
Point E (1396)
22
22
12,957
Point F (1996)
4
4
12,773
-------
TABLE 5
DERIVATION OF SECOND-ORDER RATE CONSTANTS
AND DEGRADATION RATES
Stoichiometric-
Estimated
Estimated
Measured
Measured
Measured
Normalized
Second-Order
Second-Order
Plattsburgh FT-002
BTEX
Total CAHs
Chloride
Total CAHs
Rate
Rate
Sampling Location
(ng/L)
(ng/L)
(Hg/L)
(ng/L)
(Hg/L-day)
(day'1)
MW-02-014 (1993)
6,010
10,080
63,500
7,974
2.27
0.00023
84-M (1993)
2,751
1,320
94,300
1,073
0.14
0.00010
MW-02-0I9 (1993)
2,399
3,542
42,400
2,879
0.32
0.000090
MW-02-042 (1993)
57
1,669
11,300
1,335
0.0036
0.0000021
MW-02-043 (1993)
0
384
930
191
0.0000058
0.000000015
AVERAGE of
first three points
0.5460
0.00014
Point A (1995)
16,790
76,692
63.000
56,936
6.32
0.000082
Point B (1995)
3,060
14,970
48,000
12,170
0.22
0.000015
Point C (1995)
3,543
10,038
46,000
8,160
0.17
0.000017
Point D (1995)
89
1,423
14,000
1,157
0.00062
0.00000044
Point E (1995)
40
2,242
20,000
1,817
0.00044
0.00000020
Point F (1995)
2
227
3,000
185
0.0000022
0.000000010
AVERAGE of
first four points
1.1205
0.000029
FIGURE 1
PLATTSBURGH TOTAL ATTENUATION RATE FROM 1995
DATA
100000
-5.2E-04X
y = 88433e'
R2 = 0.95
10000
1000
100
8000
12000
4000
6000
10000
2000
TRAVEL TIME (days)
-------
FIGURE 2
PLATTSBURGH BIODEGRADATION RATE ASSUMING STEADY-
STATE CONDITIONS FROM 1995 DATA
100,000
M
3
2
O
H
2
m
y
2
O
u
X
6
_9
H
O
H
10,000
1,000
-2.5E-04X
y = 88433e
R = 0.95
500
1000
1500
2000
2500
3000
3500
TRAVEL DISTANCE (feet)
FIGURE 3
PLATTSBURGH REDUCTIVE DECHLORINATION RATE FROM
1995 DATA
100000
1
3
z
2
H
2
hJ
u
2
O
u
ss
<.
u
J
<
b-
o
H
10000 -
r
-y = 77048e
•R2 = 0.98
-6.3B-05X
Aerobic Zone - No Reductive
Dechlorination
0
2000 4000 6000 8000
TRAVEL TIME (days)
10000
12000
-------
FIGURE 4
MEASURED BTEX vs. STOICHIOMETRIC CAH (1993)
0.0001
0 <~
<
o
I
I
o
<
js
a
+
o
-0.0001
-0.0002 --
-0.0003 --
-0.0004 --
-0.0005
10000
12000
y = -3.8E-08x + 8.1E-05
R = 0.92
TRAVEL TIME (days)
FIGURE 5
MEASURED BTEX vs. STOICHIOMETRIC CAH (1995)
14000
0.00002
<
o
E
a
J3
o
r—^
a
+
0.00001
-0.00001
0.00002
0.00003
-0.00004
2,000
4,000
y = -4.9E-09x + 1.1E-05
R2 = 0.81
6,000
8,000
10,000
12,300
-0.00005
TRAVEL TIME (days)
-------
TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing
1. REPORT NO, I 2.
EPA/i00M-9t/123 |
3, R£
4. TITLE AND SUBTITLE
Estimating the Changing Rate of Anaerobic Reduct-
ive Dechlorination of Chlorinated Aliphatic Hydrc
carbons in the-Presence of Petroleum Hydrocarbons
5. R6I
a. performing organization coos
7. AuTHORts) Rave e . Moutoux, Leigh Alvardo Benson,
Matthew A. Swanson,& Todd Wiedemeier (1) John T.
Wilson (2) Jerrv E. Hansen f3)John Lenhart (4)
8. PERFORMING organization report no.
3. PERFORMING ORGANIZATION NAME AND AQQRESS
(1) Parsons Engineering Science, Inc.
(2) US EPA, National Risk Management Research Lai
(3) Air Force Center for Environmental Excellence
(4) Division Environmental Science & Engineering,
10. PROGRAM ELEMENT no.
11. CONTRACT/GRANT HQ.
In-House KPDK2
Colorado School of Mines
12. SPONSORING AGENCY NAME AND AOORESS
U. S. EPA, NRMRL, SPED
P.O. Box 1198
Ada, Oklahoma 74820
13. TYPE OP REPORT ANO PCRIOO COVERED
BOOK CHAPTER
14. SPONSORING AGENCY COOS
EPA/600/15
is. supplementary notes
IS. ABSTRACT ' "" "
, R-ecent laboratory and Seid results deaaonsrate thai different chlorinated aliohatic
aydrocarocas (CaHs) ultimately caa be transformed into innocuous chemical comeouads ia
many aqiaier r/ssesu. Transformation. can be the result of either comeabolic reactions or
recucapa oxiaadoa (redox) rescdons. m the latter, chlorinated compounds sods. as
tesacmoroetheae (PCI), jrichloroetheae (TCE), and dichioroethene (DCS) caa be used is
aaaeroeis._ reaucmg aquifes as alternate electron accepters dtaiag the oxidation of organic
matter sues, as the petroleum hydrocarbons isessns, toluene, ethylbenzese, and xvienes fBTEX).
wpea usee, as electron acceptors, these CAH camoouads are secuentiailv dechlcrinaad into less
caionaaisd compounds such as viayi chloride (VQ or ethsae". White the rate of oetratesn
iycrccaroon degradation is generally reasonably approximated a a first-order srscsss, it is not
ctear tijix the degradation of elecooa acceptors such as CaK comootiads are best estimated ia a
similar sasaer, Because the rare of reductive dechlorination of CaHs ia the presence of
pejaieuzs aycsocaroons is a fcacaoa of both the concentration of the aetroletim hydrocarbons
aad me concentration of the electron accentors themselves, a second-order Mnetic'ctodei may
pro vice a oetter approxnnaden of the rate of BTEX/CaH degradation over time and distance.
Several methodologies may be used to estimate the rate of reductive dechiorinadon of CAK
compounds 'Mies they are being used to oxidize BTHX compounds. Both Srss^rder and second-
orosr approximations of CAH degradation rates can be use'fui ia predicting CAH/3TEX plume
csnavior, although the_ second-otder approximation will be especially useful ia deteammcg
waetr.er ens system will first exhaust the available supoiy of electron donors (i.e., starve) or
electron acceptors (Is., strangle).
17. KEY WORDS ANQ DOCUMENT ANALYSIS
». DESCRIPTORS
b. lOENTIPiERS/OPEN SNOED TERMS
€. COSATI Field. Group
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Release To Public
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